Complete degradation of di-n-butyl phthalate by Glutamicibacter sp. strain 0426 with a novel pathway

doi:10.21203/rs.3.rs-2696404/v1

Di-n-butyl phthalate (DBP) is widely used as plasticizer that has potential carcinogenic, teratogenic, and endocrine effects. In the present study, an efficient DBP-degrading bacterial strain 0426 was isolated and identified as a Glutamicibacter sp. Strain 0426, which can utilize DBP as the sole source of carbon and energy and completely degraded 300 mg/L of DBP within 12 hours. The optimal conditions (pH 6.9 and 31.7°C) for DBP degradation were determined by response surface methodology and DBP degradation well fitted with the first-order kinetics. Bioaugmentation of contaminated soil with strain 0426 enhanced DBP (1 mg/g soil) degradation, indicating the application potential of strain 0426 for environment DBP removal. Strain 0426 harbors a distinctive DBP hydrolysis mechanism with two parallel benzoate metabolic pathways, which may account for the remarkable performance of DBP degradation. Sequences alignment has shown that an alpha/beta fold hydrolase (WP_083586847.1) contained a conserved catalytic triad and pentapeptide motif (GX1SX2G), of which function is similar to phthalic acid ester (PAEs) hydrolases and lipases that can efficiently catalyze hydrolysis of water-insoluble substrates. Furthermore, phthalic acid was converted to benzoate by decarboxylation, which entered into two different pathways: one is the protocatechuic acid pathway under the role of pca cluster, and the other is the catechol pathway. This study demonstrates a novel DBP degradation pathway, which broadens our understanding of the mechanisms of PAE biodegradation.

Di-n-butyl phthalate

Biodegradation

Degradation pathway

Soil remediation

Genome analysis

Phthalate acid esters (PAEs) are a group of synthetic compounds widely used as additives in the manufacture of plastics and also in the production of adhesives, paints, lubricants, pharmaceutics, cosmetics, and insecticide carriers (Whangsuk et al., 2015). The global production of phthalates has increased significantly over recent decades and reached more than 8 million tons in 2011 (Gao and Wen, 2016; Gani et al., 2017). Among the various PAEs, dibutyl phthalate (DBP) is one of the most widely used phthalates, and the application and consumption of DBP has increased rapidly worldwide (Gao and Wen, 2016; Jin et al., 2012). PAEs can be easily released into the environment because they are physically (rather than chemically) bonded to the matrix of plastic products (Xu et al., 2017). Because of their potential carcinogenic, teratogenic, and endocrine effects, DBP and other PAEs have been listed as priority pollutants by both the China National Environmental Monitoring Centre and the US Environmental Protection Agency (Mo et al., 2015).

As a result of their recalcitrant chemical structure, the photo-degradation and hydrolyzation rates of these pollutants are very slow under natural conditions. Studies have demonstrated that microorganisms, mainly bacteria and fungi, play an essential role in PAE degradation under eco-friendly conditions (Sarkar et al., 2013). Bacterial strains with varying abilities in terms of degrading PAEs have been isolated from different sources such as activated sludge, contaminated soil, and wastewater (Jin et al 2012, Lu et al., 2009, Zhao et al., 2018, Mahajan et al., 2019). Whole cells of Variovorax sp. BS1 had the ability to degrade up to 600 mg/L of dimethyl phthalate (DMP) after 30 h (Prasad and Suresh, 2015). Three Pseudomonas strains degraded 71.7–74.7% of 500 mg/L di-(2-ethylhexyl) phthalate (DEHP) within 44 h (Singh et al., 2017). Gordonia sp. strain Dop5 completely degraded 750 mg/L of di-n-octyl phthalate within 40 h (Sarkar et al., 2013). Rhodococcus ruber depleted DBP with a half-life of 5.2 d in soil (Li et al., 2006). The pathway of aerobic degradation of PAEs is common to most reported microbes and is initiated by ester hydrolysis to the corresponding mono-alkyl phthalate acid ester. The monoester is then further hydrolyzed to phthalic acid, which is converted to protocatechuic acid via different dioxygenase-catalyzed pathways in Gram-positive and Gram-negative bacteria (Gao and Wen, 2016, Gani et al., 2017). The mineralization of phthalic acid by the dioxygenase-catalyzed pathway is thought to be a key step in the PAEs biodegradation process (Gao and Wen, 2016). Therefore, complete mineralization of PAEs requires a series of enzymes. Although some individual microbes possess the entire catabolic pathway, the complete degradation of PAEs is always performed by the synergy of various microorganisms in natural environments (Chatterjee and Karlovsky, 2010; Gao and Wen, 2016). Studies have reported that it usually takes days for bacteria to degrade PAEs at different concentrations, whereas degradation can be accomplished within hours for some fungal strains (Chatterjee and Karlovsky, 2010; Gao and Wen, 2016). It has been noted that the biodegradation efficiencies of phthalates under natural environmental conditions are lower than under laboratory conditions owing to the adsorption of phthalate esters onto sediments, soils, organic matrices, and other influencing factors (Gao and Wen, 2016).

To cleanup PAEs contamination in an environment, a highly effective microbial degrader is vital. In the present study, we isolated a new DBP degrader, strain 0426, which was identified as a strain of Glutamicibacter sp. The purpose of the present study was to learn the mechanism of DBP degradation and examine the application potential of strain 0426. We investigated the kinetics and factors influencing DBP biodegradation. Meanwhile, we evaluated DBP biodegradation in the soil bioaugmented with strain 0426. Further, the metabolites of DBP biodegradation and the genome were analyzed, and a completed degradation pathway was deduced and a novel molecular mechanism for DBP degradation was revealed.

Chemicals and soil

Methanol and dichloromethane of high-performance liquid chromatography grade, and DBP of analytical grade, were purchased from Sigma-Aldrich (St. Louis, MO, USA). Other chemicals were of analytical grade and were obtained from Sangon Biotech Co., Ltd. (Shanghai, China). The soil samples used in this study were collected from Jinan (N36°40′, E117°03′), and their characteristics are summarized in Table S1. The soil samples were air dried and sieved (2 mm) to remove stones and debris before used.

Culture enrichment and isolation

Carbon-free mineral medium (CFMM) was prepared as previously described by Li et al. (2006), and Luria–Bertani (LB) medium was prepared according to Sambrook et al. (1989). Bacterial strain used in this study was newly isolated from the collected soil samples by an enrichment-culture technique. The enrichment and isolation were conducted by adding soil samples into CFMM liquid medium supplemented with DBP (dissolved in methanol as a stock solution) as sole carbon and energy sources. The suspension was incubated aerobically at 30°C in the dark with vigorous shaking at 180 rpm. Aliquots (10%, v/v) were transferred to fresh medium after 5 d and incubated under the same conditions, while DBP was supplemented gradually from 50 mg/L to 300 mg/L. A pure culture was obtained by plating the enrichment culture onto CFMM solid medium containing 300 mg/L DBP.

Genome sequencing, gene analysis, and nucleotide sequence accession number

The genome of strain 0426 was sequenced on a HiSeq 4000 platform (Illumina, San Diego, CA, USA). The procedure of sequencing and analysis, including library construct, de novo genome assembly, gap closure, genome annotation and function annotation was performed as described by Chen et al. (2018). Multiple sequence alignment was conducted using Jalview software (Version 2.10.5). The whole genome sequence of Glutamicibacter sp. strain 0426 has been deposited at GenBank under accession number MPBI00000000.

Optimization of conditions for DBP degradation

To investigate the individual and interactive effects of pH and temperature on the DBP degradation by strain 0426, a response surface methodology (RSM) based on the central composite design (CCD) was conducted. Two main factors temperature and pH were chosen as independent variables and the symbols and levels of the two independent variables used in CCD are shown in Table S2. The CCD table and the response of dependent variable for DBP degradation were created using Minitab 17.3.1 (Table S3). DBP degradation experiments were conducted according to the CCD table with DBP concentration of 300 mg/L and detected the residual DBP at the 10th hour. All tests were conducted and analyzed in triplicate and were compared against controls without the inoculum. Data obtained at the 10th hour were analyzed by the response surface regression procedure of the Minitab17.3.1 software to fit the following quadratic polynomial equation (Eq. (1)):

Y = b + Σb_mX_m + Σb_mnX_mX_n + Σb_mmX_m² (1)

Where Y is the predicted response, X_m and X_n are the variables, b is the constant, b_m is the linear coefficient, b_mn is the interaction coefficient, and b_mm is the quadratic coefficient.

An analysis of variance (ANOVA) was also conducted by Minitab17.3.1 to identify the significant difference of individual and interactive effects of variables to the responses. The 95% confidence level was used to accept or reject the model used.

Degradation experiments and kinetics analysis

Biodegradation of DBP was conducted in 300-ml Erlenmeyer flasks containing 50 mL of CFMM. DBP stock solution (50 g·L^− 1) was added into the CFMM, and strain 0426 was inoculated at a concentration of 1.8×10⁷ CFU·mL^− 1. All flasks were incubated at 30°C in the dark on a reciprocal shaker at 180 rpm. Bacterial growth on DBP was verified by measuring an increase in the cell concentration (CFU·mL^− 1) of strain 0426 concomitant with a decrease in the DBP concentration. The kinetics of DBP degradation was investigated at initial concentrations of 300, 500, 700, and 1000 mg/L with an inoculum as mentioned above. All tests were conducted and analyzed in triplicate and were compared against controls without the inoculum.

The kinetics of DBP biodegradation (first-order rate constant and half-life) was determined from the linear regression of the slope on a plot of the logarithm of the DBP concentration in subsamples of the treatment versus time. The DBP degradation data collected in this study can be described by first-order kinetics model as the Eq. (2) (Ahuactzin-Pérez et al., 2014).

lnC = − kt + A (2)

Where C is the initial concentration, k is the degradation rate constant, t is the time period, and A is the constant.

The half-life can be calculated by the Eq. (3) (Ahuactzin-Pérez et al., 2014).

t_1/2 = ln2/k (3)

For all of the tests, strain 0426 was precultured and treated as described above and inoculated at a concentration of 1.8×10⁷ CFU·mL^− 1.

Analysis of DBP and its metabolites

To monitor the degradation of DBP, the residue was extracted with an equal volume of dichloromethane, and the organic phase was dried with anhydrous Na₂SO₄. The concentration of the compound was measured by gas chromatography (GC) using an Agilent 7890B equipped with an HP-5 column (length 30 m, inner diameter 0.25 mm, film thickness 0.25 µm) with a flame ionization detector operating at 300°C. The oven temperature was 290°C, and the injector temperature was 280°C.

To isolate DBP metabolites from strain 0426, CFMM cultures with 300 mg/L DBP as the sole source of carbon and energy were collected at different times. Metabolites were extracted three times from each culture by dichloromethane. The organic phases were combined and concentrated by a rotary evaporator to a suitable volume, then filtered through a 0.22-µm organic filter, dried with anhydrous Na₂SO₄, further concentrated by gentle blowing with a stream of nitrogen gas, and dissolved with methanol. The concentrated extract was analyzed by GC-mass spectrometry (GC-MS) using an Agilent GC 6890 coupled to a 5973 mass selective detector with a HP-5 MS column (length 30 m, inner diameter 0.25 mm, film thickness 0.25 µm). The metabolites were analyzed under the following conditions: helium carrier gas at a flow rate of 1.0 mL/min, and injector temperature of 280°C. The oven temperature was programmed from 80 to 290°C at a rate of 15°C/min. Metabolites were identified based on the National Institute of Standards and Technology mass spectral database.

Soil bioaugmentation experiments

The isolated strain 0426 was precultured in LB medium to exponential phase at 30 ℃, with shaking at 180 rpm. Then, cells were harvested and washed three times with CFMM. The washed cells were re-suspended in CFMM for inoculation. Twenty grams soil samples prepared above were placed into 250-mL Erlenmeyer flasks. The flasks containing soil samples were divided into two groups. One group was subjected to autoclaving at 121°C for 40 min, and the other group remained unsterilized. The DBP stock solution was mixed with water and sprayed into the soil to a final concentration of 1 mg/g soil. The moisture content was adjusted to about 15% of the soil’s water holding capacity. A cell suspension of strain 0426 was prepared as described above was added to each flask to a final cell density of around 1.8×10⁷ CFU/g soil. The flasks without strain 0426 inoculated were used as control. All tests were conducted in triplicate and incubated statically at 30°C in the dark. One gram soil sample was taken from each flask once a day to extract residual DBP by ultrasonic extraction at 260 W for 30 min with dichloromethane as extraction agent.

Isolation and identification of the DBP-degrading bacterial strain

Strain 0426 was isolated from a soil sample by enrichment culture using DBP as the sole source of carbon and energy. After sequential subculturing on CFMM agar plates supplemented with DBP, a pure culture was yielded. 16S rRNA gene sequence analysis revealed that this new DBP-degrading isolate belonged to the genus Glutamicibacter, a reclassified novel genus which previously belonged to Arthrobacter (Busse, 2016). Until now, although there have been no reports of a Glutamicibacter sp. capable of degrading PAEs, some Arthrobacter strains have been shown to degrade PAEs (Chatterjee and Dutta, 2008; Wu et al., 2010; Wen et al., 2014).

Optimization of culture conditions for DBP degradation by strain 0426

Because the two-factor interaction effects (X₁X₂) were not significant (P > 0.05) and not contained in the model (Table 1) based on the ANOVA results, the corrected quadratic polynomial equation (Eq. (4)) was adopted to describe the effects of temperature and pH on the DBP degradation rate:

Y = -712.50 + 13.08 X₁ + 173.43 X₂ − 0.2064 X₁² − 12.560X₂² (4)

Where Y is the predicted DBP degradation rate (%), X1 and X2 are the coded values for the pH, temperature, respectively.

The results of residual error diagnosis for the regression model were satisfactory (Fig. S2 A). The value of the regression coefficient (R² = 99.24%) is large enough to prove the regression model equation that can be used to explain most of the variation in the response. The adjusted R² (98.74%) and predicted coefficient R² (96.99%) both indicated the accuracy of the model. As shown in Figure. 1, the effects of temperature and pH on DBP degradation rates can be observed visually in the three dimensional response surface and contour plots. The optimal values of temperature and pH predicted by the regression model Eq. (4) were 31.7 and 6.9, respectively, and the theoretical maximum DBP degradation rate was 93.27% within 10 hours (Fig. S2 B).

Table 1

Analysis of variance (ANOVA) for the fitted quadratic polynomial model for DBP degradation.
Source	DF^a	SS^b	MS^c	F-value	P-value^*
Model	5	704.720	140.944	166.87	0.000
Linear	2	106.444	53.222	63.01	0.000
X₁	1	71.691	71.691	84.88	0.000
X₂	1	34.752	34.752	41.14	0.001
Square	2	597.121	298.560	353.47	0.000
X₁ X₁	1	67.452	67.452	79.86	0.000
X₂ X₂	1	399.642	399.642	473.15	0.000
Two-factor interaction	1	1.156	1.156	1.37	0.295
X₁ X₂	1	1.156	1.156	1.37	0.295
Error	5	4.223	0.845
Lack of fit	3	3.768	1.256	5.51	0.157
Pure error	2	0.455	0.228
Total	10	708.943

^a refers to degrees of freedom, ^b refers to sum of sequences, ^c refers to mean square.

R² = 99.40% (adjusted R² = 98.81%, predicted R² = 95.60%)

^* P＜0.05 indicates that the model terms are significant.

Biodegradation of DBP and kinetic analysis

Complete degradation of 300 mg/L DBP was achieved within 12 h companied with an increase in the cell density of strain 0426 (Fig. 2A). By comparison, the non-inoculated control showed an insignificant change in the DBP amount. Similarly, DBP at a higher initial concentration ranging from 500 to 1000 mg/L was completely removed after 20–40 h (Fig. 2B), indicating the high catabolic activity and strong virulence tolerance of strain 0426. Different bacterial strains have shown divergent catabolic abilities to transform DBP. It was reported that most strains degrade 100–1000 mg/L DBP after more than 48 h (Gao and Wen, 2016). Among these strains, Gordonia sp. strain QH-11 showed strong degradation ability removing 300–750 mg/L DBP within 45 h (Jin et al., 2012). Compared with previously reported strains, strain 0426 displayed even stronger catabolic ability for DBP degradation.

During the degradation of DBP at different concentrations by strain 0426, there was no initial lag phase detected. The exponential model represents a vigorous reaction without an initial lag. In the present study, DBP degradation by strain 0426 therefore fit well with the exponential model. First-order kinetic models were constructed for the degradation of DBP at varying initial concentrations by strain 0426. The DBP degradation rate constant and the degradation half-life are shown in Table 2. The results showed that the degradation rate decreased and the degradation half-life was prolonged by an increase in the initial DBP concentration. It has been demonstrated that microbial degradation activity for DBP decreased with an increase in its concentration, and this is related to the toxicity caused by highly concentrated DBP (Ahuactzinpérez et al., 2014).

Table 2

First order kinetic analysis of DBP degradation by *Glutamicibacter* sp. strain 0426.
Initial concentration (mg/L)	Degradation rate constant (h^− 1)	Half-life (h)	R²
300	0.2351	2.95	0.9148
500	0.1006	6.89	0.9769
700	0.0719	9.64	0.9571
1000	0.0552	12.56	0.9384

DBP biodegradation in the soil bioaugmented with strain 0426

To investigate the application potential of strain 0426, bioaugmentation of DBP-supplemented soil samples with strain 0426 was conducted. On the one hand, in the sterile soil sample bioaugmented with strain 0426, 96.6% of DBP was removed after 3 d, and 99.5% was degraded after 4 d, compared with the non-bioaugmented sterile soil sample (Fig. 3A). On the other hand, in the non-sterile soil sample, bioaugmentation with strain 0426 led to the complete removal of DBP at the 6th day (Fig. 3B). The non-sterile soil sample without bioaugmentation also showed DBP depletion; however, only 45.6% of DBP was removed after 6 d and DBP remained after this time point. Although the bioaugmented sterile-soil group removed DBP faster than the bioaugmented nonsterile-soil group in the whole experimental process, it should be noticed that the latter achieved a smaller half-time (1.10 d) than the former (1.99 d) within the first two days (Fig. 3).

Soil is thought to be a major reservoir for PAEs (Jin et al., 2015). Previous studies have shown that biodegradation of PAEs under natural environmental conditions was not as efficient as under laboratory conditions (Gao and Wen, 2016). In the present study, 99.5% of DBP (1 mg/g soil) was degraded after 4 d treatment by strain 0426, indicating the potential applications of this strain in contaminated soil. However, the degradation activity of strain 0426 was decreased in soil compared with aqueous phase tests, which might be ascribed to the adsorption of the substrate in soil organic matrices (Gao and Wen, 2016). Bioaugmentation in contaminated environments is recommended when indigenous microorganisms are not competent at cleaning up the pollutants. As shown in the present study, although indigenous bacteria can remove some of the DBP in soil, more than half of the contaminant residues remained in the soil. Bioaugmentation with strain 0426 enhanced the biodegradation process, and complete removal of DBP was achieved. During bioremediation in contaminated environments, the interaction of the inoculated microorganism with biotic and abiotic factors can influence the effect of bioaugmentation (Herrero and Stuckey, 2015), which could explain the slow-down of DBP removal after 2 days in the bioaugmented nonsterile-soil group. Therefore, the influence of biotic and abiotic factors on DBP degradation by strain 0426 in contaminated environments requires further study.

Bioinformatic analysis of the strain 0426 genome

In order to explore the genetic mechanism of PAEs catabolism by strain 0426, PAEs degrading-related genes in genome of strain 0426 were identified and compared with that of other degraders. The genome size of strain 0426 was 3.55 Mb with G + C content of 62.00%, and the annotated genome contains 3194 protein-coding sequences (CDSs), 72 RNA genes, 3 ncRNA genes, 64 tRNA genes and 47 pseudo genes (Table S4). A graphical circular genome map, which displays the general characteristic of the strain genome, was constructed (Fig. S1). The function of identified genes annotated by the NR, Swiss-Prot, GO, KEGG, and COG databases accounted for 94.2%, 59.6%, 59.4%, 46.5%, and 45.9% of the CDSs, respectively.

Based on the genome sequence, genes involved almost completed DBP degradation were identified, including many hydrolases, two gene clusters encoding for degradation of benzoic acid and protocatechuic acid. Strain 0426 contains 105 CDSs annotated as esterases or hydrolases, which may account for the strain with great potential for hydrolyzing various xenobiotics (Vamsee-Krishna and Phale, 2008). However, there were no homology of these hydrolases with dialkyl or monoalkyl PAEs hydrolases as reported, which were responsible for ester bond hydrolysis of PAEs. Protein sequences alignment among an alpha/beta fold hydrolase (WP_083586847.1) of strain 0426, two dialkyl PAEs hydrolases (AJO67804.1 and AGY55960.1), two monoalkyl PAEs hydrolases (AMJ52171.1 and BAE78500.1), and two ester hydrolases (YP_352784.1 and AAN66920.1) from different bacteria was performed. Regarding the hydrolases to be compared, the two dialkyl PAEs hydrolases can catalyze the hydrolysis of dialkyl PAEs to monoalkyl PAEs (Whangsuk et al., 2015, Jiao et al., 2013), and the two monoalkyl PAEs hydrolases catalyze monoalkyl PAEs to phthalic acid (Nahurira et al., 2018, Nishioka et al. 2006). The two ester hydrolases can concurrently catalyze the cleavage and formation of ester bond with remarkable activity and broad substrate spectrum (Liu et al. 2010). The results showed that a catalytic triad and pentapeptide motif (GX₁SX₂G) of WP_083586847.1 are well conserved (Fig. 4). The activities of a wide variety of ester hydrolases and typical lipases rely mainly on the catalytic triad usually formed by Ser, His and Asp residues, which is functionally (but not structurally) identical with that of trypsin and subtilisin (Arpigny et al. 1999). It was indicated that the function of WP_083586847.1 in strain 0426 was similar to that of typical lipases and other ester hydrolases, which could hydrolyze peptide linkage and ester linkage (Ahmad et al. 1995, Arpigny et al. 1999, Mahajan et al., 2019). Additionally, lipolytic enzymes are included into alpha/beta hydrolases, and the true lipases possess peculiar catalytic properties that they can markedly catalyze water-insoluble substrates in heterogeneous system and they can adsorb on the oil/water interface before hydrolysis by the changes of the enzyme's architecture during catalysis (Arpigny et al. 1999). Coincidentally, this hydrolase was encoded by a gene within pca cluster and located between pcaB and pcaL (Fig. 5). All the results may account for the marked performance of DBP hydrolysis and the corresponding intermediates by hydrolysis.

The pca cluster is involved in the degradation of protocatechuic acid, which is a central intermediate during aerobic degradation of PAEs and polycyclic aromatic hydrocarbons by bacteria (Fan et al., 2018). Similar to pca clusters of other strains, pcaGH-pcaB-pcaL-pcaIJ was present in the pca cluster of strain 0426 (Fig. 5). However, pcaF was absent but a benzoate transporter gene located on the 899 bp upstream of pcaL and the gene encoding 4-hydroxybenzoate 3-monooxygenase followed behind pcaH (Fig. 5). 4-hydroxybenzoate 3-monooxygenase is the enzyme that catalyzes 4-hydroxybenzoate to protocatechuic acid. This may indicate a pathway of DBP degradation from phthalic acid to protocatechuic acid different from the general pathway, which depends on the pht cluster as reported in Mycobacterium vanbaalenii PYR-1(Stingley et al., 2004), Terrabacter sp. DBF63 (Habe et al., 2003), Arthrobacter keyseri 12B (Eaton, 2001), and Gordonia sp. HS-NH1 (Li et al., 2016). The genome of strain 0426 contains all genes responsible for complete degradation of PAEs, which may expand its environments bioremediation potential, and are conducive to reveal the mechanism of PAEs metabolism.

Metabolites, putative enzymes, and the deduced pathway of DBP degradation

During DBP degradation by strain 0426, the intermediates generated were monitored by GC/MS at different time points. The mass spectra for the main metabolites are shown in Fig. S3, and correspond to phthalic acid (Fig. S3 A), butyl methyl phthalate (BMP, Fig. S3 B), ethyl methyl phthalate (EMP, Fig. S3 C), dimethyl phthalate (DMP, Fig. S3 D), methyl phthalate (Fig. S3 E), and butanoic acid (Fig. S3 F).

Based on the identified metabolites and the putative enzymes involved in degradation revealed by genome sequence analysis, a complete DBP degradation pathway was deduced in strain 0426 (Fig. 6). In the proposed pathway, DBP was converted to phthalic acid via hydrolysis, besides phthalic acid, BMP, EMP and DMP were formed. This process was different from the PAEs initial hydrolysis occurred in other bacteria that DBP was firstly hydrolyzed into mono-n-butyl phthalate and further hydrolyzed into phthalic acid (Jin et al., 2012; Huang et al., 2018; Mahajan et al., 2019; Wu et al., 2013; Whangsuk et al., 2015). It was indicated that there may be a special hydrolysis mechanism of DBP in strain 0426, which related to the peculiar hydrolyases harbored by strain 0426 as described above. Particularly, the hydrolase WP_083586847.1 may catalyze a reverse reaction of hydrolysis and esterification of mono-n-butyl phthalate to BMP, and phthalic acid to EMP/DMP. Phthalic acid was converted to benzoate by decarboxylation and then formed 4-hydroxybenzoate under the role of benzoate 4-monooxygenase. Protocatechuic acid was produced from 4-hydroxybenzoate by the catalysis of 4-hydroxybenzoate 3-monooxygenase (WP_073707416.1). Under the role of pca cluster, protocatechuic acid underwent successive oxygenolytic ring-cleavage, rearomatization, decarboxylation, and hydrolysis, and transformed into 3-ketoadipayl-CoA, which finally transformed into succinyl-CoA and acetyl-CoA, and entered into the TCA cycle. Although protocatechuate, 3-carboxy-cis, cis-muconate, 4-carboxy muconolactone, 3-ketoadipate enol lactone, 3-ketoadipate, and 3-ketoadipayl-CoA were not detected by GC-MS, the presence of the coding genes of the enzymes responsible for their transformation was revealed by analysis of the genome sequence of strain 0426. As to the existence of benzoic acid degradation gene cluster, benzoic acid in the proposed pathway might branch into cis-1,2-dihydroxycyclohexa-3,5-diene-1-carboxylate pathway, followed by the formation of catechol, cis, cis-muconate, and (+)-muconolactone, then joined into the 3-ketoadipate enol-lactone degradation.

In Gram-positive bacteria, phthalic acid is oxygenated to form 3,4-dihydro-3,4-dihydroxyphthalate, then dehydrogenized to 3,4-dihydroxyphthalate and finally decarboxylated to protocatechuate (Eaton, 2001; Habe et al., 2003; Stingley et al 2004; Li et al 2016). In Gram-negative bacteria, the process will undergo oxygenation and dehydrogenation at 4 and 5 position of carbons to form 4,5-dihydroxyphthalate, followed by decarboxylation to form protocatechuate (Chang and Zylstra, 1998; Nomura et al., 1992). In Rhodococcus sp. 2G, phthalic acid transformation is divided into two pathways, one is phthalic acid progresses through decarboxylation to form benzoic acid, and the other is phthalic acid conversion into protocatechuic acid like common Gram-negative bacteria. However, phthalic acid in strain 0426 was decarboxylated to form benzoic acid firstly, then twice hydroxylated to form 4-hydroxybenzoate and protocatechuic acid, which is a novel pathway for DBP degradation.

A newly isolated bacterium Glutamicibacter sp. strain 0426 capable of utilizing DBP as the sole source of carbon and energy was characterized, which completely degraded 300 mg/L of DBP within 12 h. The optimal conditions for DBP biodegradation were pH 6.9 and 31.7°C based on the results of RSM. DBP degradation by strain 0426 followed an exponential model and the first-order degradation half-life ranged from 2.95 to 12.56 h for 300–1000 mg/L DBP. Bioaugmentation of DBP-contaminated soil with strain 0426 enhanced DBP degradation, which demonstrated the environment remediation potential of strain 0426. Based on the identified metabolites and bioinformatic analyses of putative genes involved in degradation, a complete DBP-degradation pathway was deduced. The special enzymes for DBP hydrolyzation and the novel pathway of DBP degradation broaden our understanding of PAEs degradation mechanism.

Acknowledgments

Funding was provided by the National Natural Science Foundation of China (grant no. 21876100).

Conflict of interest The authors have no competing interests to declare relevant to this article’s content.

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No competing interests reported.

SI.docx

Complete degradation of di-n-butyl phthalate by Glutamicibacter sp. strain 0426 with a novel pathway

Status:

Journal Publication

Version 1

Abstract

Figures

Introduction

Materials And Methods

Chemicals and soil

Culture enrichment and isolation

Genome sequencing, gene analysis, and nucleotide sequence accession number

Optimization of conditions for DBP degradation

Degradation experiments and kinetics analysis

Analysis of DBP and its metabolites

Soil bioaugmentation experiments

Results And Discussion

Isolation and identification of the DBP-degrading bacterial strain

Optimization of culture conditions for DBP degradation by strain 0426

Y = -712.50 + 13.08 X₁ + 173.43 X₂ − 0.2064 X₁² − 12.560X₂² (4)

Biodegradation of DBP and kinetic analysis

DBP biodegradation in the soil bioaugmented with strain 0426

Bioinformatic analysis of the strain 0426 genome

Metabolites, putative enzymes, and the deduced pathway of DBP degradation

Conclusion

Declarations

References

Additional Declarations

Supplementary Files

Status:

Journal Publication

Version 1

Complete degradation of di-n-butyl phthalate by Glutamicibacter sp. strain 0426 with a novel pathway

Status:

Journal Publication

Version 1

Abstract

Figures

Introduction

Materials And Methods

Chemicals and soil

Culture enrichment and isolation

Genome sequencing, gene analysis, and nucleotide sequence accession number

Optimization of conditions for DBP degradation

Degradation experiments and kinetics analysis

Analysis of DBP and its metabolites

Soil bioaugmentation experiments

Results And Discussion

Isolation and identification of the DBP-degrading bacterial strain

Optimization of culture conditions for DBP degradation by strain 0426

Y = -712.50 + 13.08 X1 + 173.43 X2 − 0.2064 X12 − 12.560X22 (4)

Biodegradation of DBP and kinetic analysis

DBP biodegradation in the soil bioaugmented with strain 0426

Bioinformatic analysis of the strain 0426 genome

Metabolites, putative enzymes, and the deduced pathway of DBP degradation

Conclusion

Declarations

References

Additional Declarations

Supplementary Files

Status:

Journal Publication

Version 1

Y = -712.50 + 13.08 X₁ + 173.43 X₂ − 0.2064 X₁² − 12.560X₂² (4)