Surfactin for enhanced removal of aromatic hydrocarbons during biodegradation of crude oil

Fuel 267 (2020) 117272

Contents lists available at ScienceDirect

Fuel journal homepage: www.elsevier.com/locate/fuel

Full Length Article

Surfactin for enhanced removal of aromatic hydrocarbons during biodegradation of crude oil

T

Xinwei Wanga,b, ,1, Ting Caic,1, Weitao Wend, Jiazhen Aia, Jiayi Aic,e, Zhihuan Zhangc, Lei Zhuc, Simon C. Georgee ⁎

a

College of Chemical Engineering and Environment, China University of Petroleum, Beijing 102249, China State Key Laboratory of Heavy Oil Processing, China University of Petroleum, Beijing 102249, China c State Key Laboratory of Petroleum Resources and Prospecting, China University of Petroleum, Beijing 102249, China d Safety, Environment, Quality Supervision and Testing Research Institute, CNPC, Chuanqing Drilling Engineering Co., Ltd, China e Department of Earth and Planetary Sciences and MQ Marine Research Centre, Macquarie University, North Ryde, Sydney, NSW 2109, Australia b

ARTICLE INFO

ABSTRACT

Keywords: Surfactin Crude oil Biodegradation Alkylphenanthrene Alkyldibenzothiophene Methyltriaromatic steroid

Oil biodegradation studies using the powerful surfactant have mainly focused on saturated hydrocarbons or parent aromatic hydrocarbons, but not specifically on the position or degree of alkyl substitution of aromatic hydrocarbons. Removal of substituted aromatic hydrocarbons remains a controversial topic in biodegradation. In this study, the effect of surfactin addition on specific alkylated aromatic isomers and saturated hydrocarbons was observed in aerobic biodegradation experiments with four species of bacteria. The relative susceptibility of the individual alkylphenanthrene, alkyldibenzothiophene and methyltriaromatic steroid isomers to biodegradation was determined by their depletion rates. The results show that different bacteria induce removal of alkylated aromatic hydrocarbons and alteration of other petroleum hydrocarbons in various ways. Bacillus amyloliquefaciens strain BC is the best for biodegradation of general petroleum hydrocarbons, because it can convert nalkanes, n-alkylcyclohexanes, isoprenoids, and aromatic hydrocarbons. Achromobacter strain J3 can degrade methylphenanthrenes, methyltriaromatic steroids, triaromatic steroids, and steranes better than other bacteria. Citrobacter sp strain J1 induces the highest degradation rate for dimethylphenanthrenes, trimethylphenanthrenes, and dimethyldibenzothiophenes, and Brucella melitensis strain J2 selectively degrades methyldibenzothiophenes. Surfactin addition increases the biodegradation rate of alkylaromatic hydrocarbons by 1.5–87.2%, and can significantly enhance the biodegradation of alkylphenanthrenes, alkyldibenzothiophenes and methyltriaromatic steroids. Alkyldibenzothiophenes can be used as markers for determining the levels of biodegradation of crude oils, and when used in conjunction with triaromatic steroids are a powerful indicator of the biodegradation of petroleum. The use of surfactin for enhancing the removal of aromatic hydrocarbons provides a wide range of applications in the environmental remediation and petroleum industry.

1. Introduction Biodegradation is part of the biogeochemical cycle and is the main removal pathway for petroleum compounds in water, sediments and soils environments. In this cycle, diverse bacteria have different effects on the degradation of petroleum compounds, causing different biogeochemical cycle speeds for the four main components of petroleum (saturated hydrocarbons, aromatic hydrocarbons, polars or resins, and asphaltenes) [1–3]. The biodegradation of aromatic hydrocarbons is highly dependent on the number of aromatic rings present in the molecule, with the rate of biodegradation decreasing with increasing number of aromatic rings [4].

Alkylated polycyclic aromatic hydrocarbons (PAHs) are abundant in petroleum, and can be biodegraded by the removal of alkyl substituents. The degree of alkylation is one of the critical factors controlling the rate of biodegradation, decreasing with increasing number of alkyl substituents [5]. The distributions of homologous series of alkylated aromatic hydrocarbons can be used for fingerprinting crude oils, although their composition can be significantly altered by the degradation of the low molecular weight species [6]. During biodegradation, various classes of compounds are degraded simultaneously but at significantly different rates [5,7,8]. The relative susceptibility of alkylnaphthalenes, alkylphenanthrenes and alkylbenzenes

Corresponding author at: Beijing Key Laboratory of Oil and Gas Pollution Control, China University of Petroleum, Beijing 102249, China. E-mail address: [email protected] (X. Wang). 1 These authors contributed equally to this work. ⁎

https://doi.org/10.1016/j.fuel.2020.117272 Received 24 September 2019; Received in revised form 9 January 2020; Accepted 31 January 2020 0016-2361/ © 2020 Elsevier Ltd. All rights reserved.

Fuel 267 (2020) 117272

X. Wang, et al.

Fig. 1. Total ion chromatograms (TIC) of the saturated hydrocarbon fractions of the crude oil from the Liaohe Oilfield, China (control; A), showing abundant nalkanes (numbered peaks) and C30 αβ hopane. Pr = pristane, Ph = phytane. (B) to (E) are TICs of the same crude oil after the biodegradation experiments with different microorganism strains. (F) to (I) are TICs of the same crude oil after the biodegradation experiments with different bacteria strains and added surfactin (SF; surfactin). See Table 1 for identification of the bacteria used in each experiment.

to microbial alteration has been well documented [7,9–11]. For example, Huang et al. [9] observed that methylphenanthrene concentrations decrease faster than that of phenanthrene during biodegradation of oil columns in the Liaohe Basin. The biodegradation of methyldibenzothiophenes in crude oil under an aerobic environment has not attracted much previous attention [8], although many studies have reported the susceptibility of methyldibenzothiophenes and dimethyldibenzothiophenes to biodegradation in oil reservoirs [8,9,11,12]. The environmental fate and potential utility of methyldibenzothiophenes (MDBTs) for bioremediation assessment has been studied after the Prestige oil spill [13–15]. Studies on the changes of alkylbenzothiophenes that occur in surface water environments are limited. Dutta and Harayama [6] found that n-alkylbenzothiophenes are almost completely biodegraded in sediment. Triaromatic steroids may be formed from monoaromatic steroids by

aromatization and loss of a methyl group [16] (Section 14.6). Because of their higher thermal stability and resistance to biodegradation, few studies have investigated the natural biodegradation of triaromatic steroids [4]. Aerobic biodegradation of aromatic steroid hydrocarbons in a laboratory was demonstrated by Wardroper et al. [8]. The C28 20Striaromatic steroid was found to yield more accurate estimates of depletion during biodegradation [13,17,18]. However, the relative biodegradation susceptibilities of hydrocarbons can vary with different physiochemical conditions and diverse bacteria which influence microbial functioning [9,11,19]. The dominant bacteria in an oil reservoir are likely to be different according to the properties of the crude oil and the location of the reservoir. Moreover, the preference in which different microorganisms utilize methyltriaromatic steroids compound may vary. 2

Fuel 267 (2020) 117272

X. Wang, et al.

Most of the observed variation in environmental polycyclic aromatic hydrocarbon (PAH) biodegradation rates comes from processes controlling the bioavailability of the compounds, and not the processes controlling uptake or biotransformation [20]. Production of biosurfactants by bacteria is considered to be an important microbial strategy that influences the bioavailability of hydrophobic chemicals [21,22]. Biosurfactant-producing bacteria commonly exist in environments such as water bodies, sediments and soils. By reducing surface and interfacial tension, biosurfactants can increase the solubility, mobility, bioavailability, and subsequent biodegradation of petroleum hydrocarbons [23–27]. Surfactin is a cyclic lipopeptide, produced by different strains of Bacillus subtilis, and is considered to be a powerful biosurfactant due to its excellent surface activity that is required to mobilize the entrapped oil [28,29]. Surfactin has a mixture of isoforms, due to variations in the carbon chain length and branching of its hydroxyl fatty acid components. The fatty acid alkyl chain of surfactin varies in length mainly from 13 to 15 carbon atoms [29]. The production, characterization, function, and application of surfactin have been previously studied [23,24,30–32]. The ability of surfactin to accelerate the biodegradation of n-alkanes and phenanthrene has been widely studied [22,23,26,33–35]. However, the effect of surfactin on the biodegradation of alkylated aromatic compounds has been little studied. Much work is needed to investigate the ability of microorganisms to efficiently degrade crude oil, and for them to produce eco-friendly biological surfactants with promising environmental applications. This work focused on the effect of surfactin on alkylated aromatic compounds, with the aim of exploring the relative susceptibility of the individual alkylphenanthrenes, alkyldibenzothiophenes and methyltriaromatic steroids to biodegradation. The effect of surfactin on petroleum hydrocarbons was also studied, so as to assess the contribution of surfactin to the removal of methylaromatic hydrocarbons by aerobic bacteria in the surface environment. Characterization of the surfactin along with its capacity to remove alkylated aromatic isomers provides a wide range of applications for the remediation of environments contaminated by crude oil.

Fig. 2. Molecular structure of surfactin.

2.3. Microorganisms and biodegradation tests Bacterial strains of Citrobacter sp, Brucella melitensis, and Achromobacter were isolated from the oil contaminated soil and the culture medium in which resins were used as the sole carbon source. Bacillus amyloliquefaciens was screened from crude oil from D212XingH290 well in the Shuguang oil production plant of the Liaohe Oilfield, China. Nine sample groups were used to explore the effects of microorganism strain and surfactin on oil biodegradation. Groups J1, J2, J3, and BC contained Citrobacter sp, Brucella melitensis, Achromobacter, and Bacillus amyloliquefaciens respectively (Table 1). Groups J1 + SF, J2 + SF, J3 + SF, and BC + SF each had a strain supplemented with surfactin at a concentration of 100 mg/L. The group control tested an uninoculated flask (Table 1). Before the biodegradation tests, an enrichment medium was made which contained 3.0 g of beef extract, 10.0 g of peptone, and 5.0 g of NaCl per litre. The four microorganism strains were cultivated in 100 mL of the enrichment medium for 24 h. Then 1 mL of each was inoculated in a flask containing 150 mL of a biodegradation media. The biodegradation media contained yeast extract (1.5 g/L), NaNO3 (1.5 g/ L), KCl (0.5 g/L), KH2PO4 (1.0 g/L) and MgSO4 (0.5 g/L). 2 g of the crude oil was used as the sole carbon source. The pH of the medium was

2. Materials and methods 2.1. Characteristics of the crude oil The crude oil sample was obtained from the Du212-XingH290 well in the Shuguang oil production plant of the Liaohe Oilfield, China. The viscosity of the oil is 15,900 mPa·s at 40 °C and the American Petroleum Institute (API) gravity of the oil is 19. The content of nitrogen, carbon, hydrogen and oxygen are 0.44, 81.4, 10.3, and 3.4 wt%, respectively. The relative abundance of saturated hydrocarbons in the oil is high. Molecular analysis by gas chromatography-mass spectrometry (GC–MS) indicates the oil contains abundant n-alkanes (Fig. 1A) and low molecular weight hydrocarbons, and no unresolved complex mixture, confirming that it is not biodegraded.

Table 1 Laboratory design for aerobic biodegradation of crude oil. Sample Code

2.2. Characteristics of surfactin The surfactin used in this study was produced by Bacillus amyloliquefaciens. Surfactin was extracted from the culture medium and purified using modified acid precipitation and solvent extraction methods [36]. Production and characteristics of surfactin have been described by Cai et al. [37]. In brief, surfactin is a cyclic lipopeptide composed of beta-hydroxy fatty acids and seven amino acids (Fig. 2). The fatty acids are normal, isomeric or anti-isomeric fatty acids with carbon chain lengths of 7–10. The sequence of amino acids is usually Glu-Leu-LeuVal-Asp-Leu-Leu. The critical micelle concentration of surfactin is 100 mg/L.

Control J1 J2 J3 BC J1 + SF J2 + SF J3 + SF BC + SF

Conditions Bacteria

Additive

No Bacteria Citrobacter sp Brucella melitensis Achromobacter Bacillus amyloliquefaciens Citrobacter sp Brucella melitensis Achromobacter Bacillus amyloliquefaciens

— — — — — Surfactin Surfactin Surfactin Surfactin

Abbreviations: SF = surfactin. 3

Fuel 267 (2020) 117272

X. Wang, et al.

adjusted to 7.0 by addition of NaOH. The culture period was 45 days at 30℃, with a speed of 160 rpm. After cultivation, the residual oil was recovered by liquid–liquid extraction in a separating funnel using 200 mL dichloromethane as the solvent (three repeats), then evaporated using a rotary evaporator. The recovered oil sample was deasphaltened using addition of an excess of n-hexane, and then fractionated using column chromatography (silica gel:alumina, 3:2 v/v) into saturated hydrocarbons using n-hexane as the eluent, aromatic hydrocarbons using dichloromethane/n-hexane (2:1, v/v), and polar compounds using dichloromethane/methanol (98:2, v/v). This separation procedure resulted in recoveries ranging from 93% to 108% (average 98%). Biodegradation rates were determined gravimetrically using Eq. (1) [21]:

Biodegradationpercent = 1

Ws × 100% W0

phytane (Ph). Samples J1 + SF, J3 + SF and BC + SF contain residual amounts of n-alkanes. The amount of biodegradation of n-alkanes for sample BC is > 95%, if C30 hopane is considered to be a conservative marker [39] (Fig. 1). It is well documented that n-alkanes can be converted to n-alcohols and then to n-fatty acids via β-oxidation during the biodegradation process [40]. In our previous resins fraction study, the relative content of the fatty acids did increase in samples J1 and J3 [21], which are likely to be the biodegradation products of n-alkanes. It’s worth noting that, sample BC was more depleted in n-alkanes than that in BC + SF. The reason for that was the surfactin was produced by strain BC. The strain BC could continuously product surfactin during biodegradation process. The concentration of surfactin which the groups supplemented was 100 mg/L, just its critical micelle concentration. Above the CMC, surfactant can detrimentally affect biodegradation. Micelle cores can trap organic contaminants, creating a barrier between microorganisms and organic molecules, resulting in the potential substrate becoming less rather than more available [25]. In addition, the adsorption of biosurfactant on microorganisms caused the cell surface hydrophobicity was depended on the physiological status of cells and the concentration of biosurfactant [27]. The excessive surfactin in sample BC + SF hindered the contraction of cell and n-alkanes, resulting in less depleted in n-alkanes than that in sample BC. Although the relative proportions of Pr and Ph are governed by complex processes [41,42], aerobic bacteria are known to preferentially consume n-alkanes, followed later during longer or more extensive biodegradation by branched alkanes such as these isoprenoids [4]. Therefore, the Pr/Ph, Pr/n-C17 and Ph/n-C18 ratios (Table 2) can be used to indicate the extent of biodegradation [10]. Pr and Ph were partially removed from most samples during the biodegradation experiment (Fig. 1), but are predominantly unaltered in sample J2. However, sample J2 + SF is strongly biodegraded (Figs. 1 and 3a). The Pr/Ph ratio of the control is 0.67 (Table 2), but is significantly reduced in samples J1 + SF, J2 + SF, J3 + SF, BC + SF and BC, demonstrating that in these samples, Pr was selectively and partly removed during biodegradation, compared to Ph. This finding is consistent with Hao and Lu [43] and cited papers therein, who showed that Pr is more susceptible to biodegradation than Ph. The low Pr/n-C17 and Ph/n-C18 ratios that are similar to the control for samples J1 and J2 without surfactin (< 1, Table 2) are indicative of low levels of biodegradation [44]. In contrast, most samples with surfactin have higher Pr/n-C17 and Ph/n-C18 ratios than the corresponding experiment without surfactin, indicating that the addition of surfactin has increased the biodegradation of the crude oil (Table 2). The exceptions are samples J3 and J3 + SF, which have very similar Pr/n-C17 and Ph/n-C18 ratios, indicating that addition of surfactin to Achromobacter did not speed up nalkane removal. The Pr/n-C17 and Ph/n-C18 ratios for sample J2 + SF are extremely high (Fig. 3a; Table 2), showing that this sample experienced the greatest n-alkane loss, consistent with moderately heavily biodegraded crude oils [10]. n-Alkylcyclohexanes are abundant in the crude oil control (Fig. 3b), and were variable removed by the biodegradation. n-Alkylcyclohexanes were almost completely depleted in all samples containing added surfactin, especially sample J2 + SF, and were also significantly removed from the BC sample. In contrast, the considerable abundance of n-alkylcyclohexanes in the other samples exposed to bacteria without surfactin (J1, J2, J3; Fig. 3b) is consistent with the cyclic alkanes having significant resistance to microbial degradation, as has previously been demonstrated [5]. The distribution patterns of C27, C28, and C29 regular steranes and diasteranes in the control and biodegraded samples are shown by the partial m/z 217 mass chromatograms (Fig. 4). There are no dramatic differences in the distributions of regular steranes for samples J1, J2, J3 and BC relative to the control sample, which are characterized by a high abundance of C29 and moderate abundances of C27 and C28 regular steranes (Table 2). However, there are slight differences in three of the samples with surfactin (J1 + SF, J3 + SF and BC + SF), which are

(1)

where W0 is the content of the oil or specific compounds in oil from the control, and Ws is the content of the oil or specific compounds in oil from the treated group after biodegradation. 2.4. Gas chromatography-mass spectrometry analysis Saturated and aromatic hydrocarbons were analysed by GC–MS (Agilent 6890/5975; USA). Helium was used as the carrier gas. The oven temperature programme for the saturated hydrocarbons was 50–120 °C at 20 °C/min, and then to 310 °C at 3 °C/min. The oven temperature programme for the aromatic hydrocarbons was a hold at 80° for 1 min, then an increase to 300 °C at 3 °C/min, and a hold for 18 min. The injector temperature was held at 300 °C and was operated in splitless mode. The injection volume was 1.0 µL from a 10 µL syringe. The MS was operated in electron impact ionization (EI) mode at 70 eV electron energy, and a scanning range of 50–600 Da. Tetracosane-D50 and terphenyl-D14 were added to the fractions as internal standards, prior to GC–MS analysis. Aromatic hydrocarbons were identified by comparison with published retention time data. For example, trimethylphenanthrenes were identified by comparison with Huang et al. [9]. 3. Results and discussion 3.1. The effect of surfactin on the biodegradation of saturated hydrocarbons The total ion chromatograms (TIC) of the saturated hydrocarbon fractions show that the most abundant hydrocarbon is always 17α(H),21β(H)-C30 hopane (Fig. 1). The n-alkanes have a unimodal distribution from n-C14 to n-C35. Dominance of odd-over-even n-alkanes (from n-C23 to n-C27) was observed in the control and the biodegraded sample J3. Compared to the control, there is depletion of the n-alkanes relative to the isoprenoids and C30 hopane in the biodegraded samples (Fig. 1). This is most marked in the Bacillus amyloliquefaciens sample (BC) and the four biodegraded samples that also contain surfactin (J1 + SF, J2 + SF, J3 + SF and BC + SF). The high molecular weight peaks in the TICs that become very obvious in the four biodegraded samples that also contain surfactin are mainly due to hopanes and steranes. The effect of biodegradation on the n-alkanes and isoprenoids is more clearly shown by the m/z 85 mass chromatograms (Fig. 3a). In general, n-alkanes are the first to be removed during biodegradation [5], whereas isoprenoids have intermediate stability, and hopanes are the most resistance of these hydrocarbons [38]. Major differences are apparent between samples with surfactin and samples without surfactin (Fig. 1). Compared to the control sample, all samples with added surfactin are strongly depleted in n-alkanes (Figs. 1 and 3a), especially sample J2 + SF in which there was complete loss of n-alkanes between n-C14 and n-C35, and partial loss of the isoprenoids pristane (Pr) and 4

Fuel 267 (2020) 117272

X. Wang, et al.

Fig. 3. Partial mass chromatograms (m/z 85 and m/z 82) of the saturated hydrocarbon fraction of the crude oil control and crude oil samples after biodegradation, showing (a) n-alkanes and isoprenoids, and (b) n-alkylcyclohexanes. Numbered peaks in (a) are n-alkanes, and in (b) are n-alkylcyclohexanes, x = alkylcyclohexane; Pr = pristane, Ph = phytane. Peaks labelled small number in (b) are n-alkanes. See Table 1 for identification of the bacteria used in each experiment.

altered by the biodegradation experienced in this study. The Ts/Tm ratio under the action of surfactin is lower than that without surfactin in 3 of the 4 experiments (Table 2), suggesting that surfactin may accelerate the degradation of Ts, which is not consistent with a previous laboratory study of aerobic biodegradation [45]. Compared with other saturated compound groups, hopanes and tricyclic terpanes are more resistant to biodegradation [5]. It is clear from the above results that biodegradation was quite significant in all the samples, and that degradation efficiency for each bacterium was different (Fig. 1). Bacillus amyloliquefaciens strain BC is the best for biodegradation of petroleum hydrocarbons, as it can convert n-alkanes, n-alkylcyclohexanes and isoprenoids. Furthermore, biodegradation degree was higher for the crude oil experiments with

characterized by lower relative amounts of C27 regular steranes. This implies that the order of susceptibility to biodegradation in this study is C27 > C28 > C29 for the regular steranes, which is consistent with previous studies [8]. The biodegraded crude oil samples have near identical C27 βα diasterane/C29 ααα sterane, C29 ααα sterane [20S/ (20S + 20R)] and C29 sterane [αββ/(αββ + ααα)] ratios to the crude oil control (Table 2), indicating that at the extent of biodegradation in this study, there is no biodegradation selectivity for the sterane isomers or for steranes relative to diasteranes. The distribution patterns of tricyclic terpanes and hopanes in the control and biodegraded samples are shown by the partial m/z 191 mass chromatograms (Fig. 5). There is little difference between the samples, indicating that for the most part these hydrocarbons were not 5

Fuel 267 (2020) 117272

X. Wang, et al.

Table 2 Ratios based on saturated hydrocarbons for the crude oil control and the crude oil samples after biodegradation. See Table 1 for identification of the bacteria used in each experiment. Ratio

Control

J1

J1 + SF

J2

J2 + SF

J3

J3 + SF

BC

BC + SF

Pr/Ph Pr/n-C17 Ph/n-C18 Pr/C30Hopane Ph/C30Hopane Ts/Tm C27:C28:C29 ααα 20R steranes C27 βα diasteranes/C29 ααα steranes C29 ααα steranes [20S/(20S + 20R)] C29 steranes [αββ/(αββ + ααα)]

0.67 0.57 0.85 0.75 1.12 0.51 0.63:0.51:1 0.05 0.34 0.31

0.70 0.57 0.87 0.82 1.17 0.52 0.63:0.51:1 0.05 0.34 0.31

0.61 1.19 1.90 0.15 0.25 0.49 0.60:0.51:1 0.04 0.34 0.31

0.70 0.62 0.93 0.77 1.11 0.52 0.61:0.51:1 0.05 0.34 0.30

0.57 27.1 61.3 0.30 0.53 0.46 0.63:0.52:1 0.05 0.34 0.31

0.69 0.82 1.19 0.65 0.95 0.51 0.58:0.50:1 0.05 0.33 0.31

0.62 0.81 1.19 0.06 0.09 0.48 0.54:0.61:1 0.06 0.38 0.34

0.58 2.50 3.80 0.10 0.17 0.50 0.63:0.50:1 0.05 0.33 0.28

0.60 1.20 1.90 0.10 0.17 0.48 0.58:0.54:1 0.05 0.35 0.31

Abbreviations: Pr = pristane; Ph = phytane; Ts/Tm = 18α-trisnorneohopane/17α-trisnorhopane); C27 βα diasteranes/C29 ααα steranes = [13β,17α 20S + 20R] diasteranes/[5α,14α,17α-24-ethylcholestanes 20S + 20R]; C29 ααα steranes [20S/(20S + 20R)] = ααα 24-ethylcholestanes [20S/(20S + 20R)]; C29 steranes [αββ/(αββ + ααα)] = [αββ 24-ethylcholestanes (20S + 20R)]/[αββ 24-ethylcholestanes (20S + 20R) + ααα 24-ethylcholestanes (20S + 20R)].

surfactin than without, showing that the addition of surfactin accelerates the biodegradation process. The process by which surfactin aids biodegradation is well known. Lipoproteins can increase the aqueous solubilization of poorly soluble compounds, making them more bioavailable, and can aid emulsification, change cell hydrophobicity, and decrease the surface tension, all of which enhance hydrocarbon bioremediation [23,24,26,28,30–32]. Since surfactin has two acidic amino acid residues (Asp and Glu), it can take non-dissociate, mono-dissociate, and di-dissociate forms, depending on the pH of the medium. Asp and Glu are known to influence the cell surface and biological functions [27]. Surfactin can also gradually penetrate into the outer layer of the phospholipid bilayer membrane [46]. Some bacteria show a similar ability to enhance biodegradation. For example, Pseudomonas poae BA1, Acinetobacter bouvetii BP18, Bacillus thuringiensis BG3, Stenotrophomonas rhizophila BG32, and Pseudomonas Pseudomonas sp. WJ6 can convert aliphatic hydrocarbons, branched alkanes, cyclic alkanes and aromatic hydrocarbons due to their production of rhamnolipids and lipopeptides [22,31]. Lai et al.[35] showed that surfactin can remove total petroleum hydrocarbons from soil. Lipoprotein-producing strains of bacteria show a 16–28% increase in ability to degrade hydrocarbons [26].

3-MP were depleted faster than 9-MP and 1-MP, and are therefore more susceptible to microbial attack. Aromatic hydrocarbons containing adjacent unsubstituted α and β positions are more readily degraded, thus 2-MP and 3-MP were previously shown to be less resistant to degradation than other methylphenanthrenes [47]. The new result is largely consistent with field observations that found that 2-MP is less resistant to biodegradation and 9-MP and 1-MP isomers have higher resistance [10,15]. Studies of naturally biodegraded crude oils indicate that 9-MP is more resistant to biodegradation than other methylphenanthrene isomers [9,48]. For example, the 2-MP/9-MP ratio has been used to evaluate the bioremediation of hydrocarbon-contaminated soils [19]. However, some studies have reported different results. The order of susceptibility to microbial attack of methylphenanthrenes in hydrocarbon-contaminated coastal sediments was 2-MP > 9-MP > 1MP = 3-MP [7]. For reservoired oils from the Panyu Oil Field, the order of susceptibility to biodegradation was 3-MP > 2-MP > 1-MP > 9MP [10]. The aerobic biodegradation of Arabian light crude oil in the laboratory resulted in an order of susceptibility of 3-MP > 2-MP = 9MP > 1-MP [49]. A preferential alteration of 1-MP has been observed in some Australian biodegraded oils and in laboratory biodegradation experiments [48,50]. Contributing to this complexity are the two simultaneous metabolic pathways for removal of 2-MP, oxidation of aromatic rings and methyl group oxidation [51,52]. In contrast, the only pathway of 9-MP removal is by dihydroxylation of the aromatic ring, because oxidation of the methyl group at position 9 is hindered by the proximity of the ring [52]. The depletion of the dimethylphenanthrenes (Fig. 8) is similar to that of the trimethylphenanthrenes (Fig. 9), but there was a lesser enhancement effect by surfactin addition compared to the methylphenanthrenes. Surfactin can enhance biodegradation of dimethylphenanthrenes by 3.1–39.0%, and trimethylphenanthrenes by 3.9–27.9%. The complexity of the GC profiles of the dimethylphenanthrenes, with many overlapping peaks due to the large number of possible isomers, means that it is not possible to determine the susceptibility to microbial attack for most individual compounds. Except for co-eluting hydrocarbons, the depletion order of dimethylphenanthrenes is 1,2-DMP > 1,8-DMP > 2,3-DMP = 1,7-DMP (Fig. 8). Several studies have shown that isomers with β-substituents (e.g., 2- and 3-methyl) or with βα-positions unoccupied are more easily co-oxidized than those with adjacent methyl groups [15,48]. Aerobic biodegradation experiments have indicated that the most easily degraded dimethylphenanthrene isomer is 2,6DMP, and the most recalcitrant is 1,7-DMP [49,53]. In laboratory biodegradation experiments on heavy fuel oil, it was shown that 1,7-DMP and 2,7-DMP are the most refractory of the dimethylphenanthrenes to biodegradation, whereas 3,6-DMP, 2,6-DMP, and 2,3-DMP exhibit significant depletion [15]. The overall susceptibility of the dimethylphenanthrenes (Table 4) is consistent with previous observations that

3.2. The effect of surfactin on the removal of methylaromatic hydrocarbons The TICs of the aromatic hydrocarbon fraction show that alkylnaphthalenes and alkylphenanthrenes are the most abundant aromatic hydrocarbons in the crude oil samples, followed by triaromatic steroids (Fig. 6). The most abundant peak in all the samples is due to co-eluting 2,10- + 1,3- + 3,10- + 3,9-dimethylphenanthrene. In order to understand the effect of biosurfactants and specifically surfactin on the biodegradation of individual aromatic hydrocarbons, the depletion of the aromatic hydrocarbons was investigated in comparison with the control sample (Figs. 7–13). Due to their high volatility, the alkylnaphthalene series is not further discussed here. Depletion of phenanthrene and the methylphenanthrenes relative to the control is higher in the samples with surfactin than in the samples without (Fig. 7). Phenanthrene is mostly more degradable than the methylphenanthrenes, thus influencing the methylphenanthrene index (MPI1; Table 3). For the samples with only bacteria, the range of biodegradation rates for methylphenanthrenes is quite variable, from 2.9 to 21.1%, and is lowest for the sample exposed to Bacillus amyloliquefaciens (BC). The percentage enhancement of biodegradation with addition of surfactin is very variable between the different bacteria (12.4% to 87.2%), indicating that there may be different modes of action between surfactin and bacteria. For example, 2-MP is depleted by 33.8% and 98.5% in sample J2 + SF and J3 + SF, respectively. The order of percentage depletions for the methylphenanthrenes is 2MP > 3-MP > 1-MP > 9-MP. Lower (3-MP + 2-MP)/(9-MP + 1MP) ratios in samples with surfactin (Table 3) also show that 2-MP and 6

Fuel 267 (2020) 117272

X. Wang, et al.

Fig. 4. Partial mass chromatograms (m/z 217) of the saturated hydrocarbon fraction of the crude oil control and crude oil samples after biodegradation, showing the distribution of steranes and diasteranes. For peak identities see Table S1 of the Supplementary Information. See Table 1 for identification of the bacteria used in each experiment.

isomers with substituents at the 9 or 10 position (the K-region) are more resistant to biodegradation than isomers with substituents at other positions [9,47,48]. The lack of enzymatic attack at the K-region of phenanthrene may be a general feature, or there may be preferential cooxidation of methyl substituents situated in the A ring [47]. However, these suggestions are in contrast to observations of the biodegradation of spilled oil in costal sediments reported by Fisher et al. [7], and in reservoir oils by Huang et al. [9], who found that 1,7-DMP is more susceptible to biodegradation than other C2-P isomers. From this work (Fig. 9), a tentative order of susceptibility to biodegradation for the trimethylphenanthrenes (from most susceptible to least; only for the non-co-eluting isomers) is inferred to be: 1,2,6TMP > 2,3,10-TMP = 1,3,8-TMP > 1,2,8-TMP = 1,6,7-TMP. The

isomer 1,2,6-TMP was more rapidly depleted than the other trimethylphenanthrenes, and is therefore more susceptible to microbial attack. The moderate sensitivity to biodegradation of 1,2,8-TMP is consistent with the finding of Permanyer et al. [54], although in reservoired oils from the Liaohe basin 1,2,8-TMP was shown to be the least-resistant trimethylphenanthrene to biodegradation [9]. These differences in the biodegradation susceptibility of alkylphenanthrenes are likely due to specific environmental conditions, such as different types of bacteria, temperature differences, or oxygen availability [19]. Different microorganisms in a mixed culture probably follow different degradation pathways [49], and the various molecular species are thus biodegraded at different rates [54]. Thus the differences may be at least partly due to the different bacterial consortium in 7

Fuel 267 (2020) 117272

X. Wang, et al.

Fig. 5. Partial mass chromatograms (m/z 191) of the saturated hydrocarbon fraction of the crude oil control and crude oil samples after biodegradation, showing the distribution of tricyclic terpanes and hopanes. For peak identities see Table S2 of the Supplementary Information. See Table 1 for identification of the bacteria used in each experiment.

different laboratory experiments and field observations [10]. Individual members of the degrader community have preferences for different alkylphenanthrene isomers during environmental biodegradation [20]. In this study, Achromobacter strain J3 can degrade methylphenanthrenes more rapidly than the other bacteria (Fig. 7), whereas Citrobacter sp strain J1 shows the highest degradation rate for dimethylphenanthrenes and trimethylphenanthrenes (Figs. 8 and 9). The enhancing effect of surfactin on the biodegradation of alkylphenanthrenes is obvious in the studied samples (Figs. 7–9). Similar results were also obtained in many previous studies, which found that lipopeptides significantly enhanced the degradation of phenanthrene [22,23,26,33]. Lipopeptides can also contribute to the emulsifying of different kinds of oil, especially crude oils [26].

The biodegradation characteristics of dibenzothiophene and the methyldibenzothiophene isomers is illustrated in Fig. 10. The biodegradation rate of the methyldibenzothiophenes is 5.0–34.4%, and the biodegradation rate was increased by 11.0–74.5% by the addition of surfactin. The behaviour of dibenzothiophene is similar to that of the methyldibenzothiophenes (Fig. 10). The order of susceptibility of the methyldibenzothiophenes to biodegradation (from most susceptible to least) is 2-MDBT + 3-MDBT > 1-MDBT > 4-MDBT, which is in good agreement with previous results [47,49]. The selective biodegradation of the methyldibenzothiophenes occurs because isomers with methyl substituents in the β-position (2-MDBT and 3-MDBT) are easily degraded [11,15,47,55]. Oxidation products of methyldibenzothiophene isomers also show that the unsubstituted ring was preferentially 8

Fuel 267 (2020) 117272

X. Wang, et al.

Fig. 6. Total ion chromatograms (TIC) of the of aromatic hydrocarbon fractions of the crude oil from the Liaohe Oilfield, China (control; J), showing abundant alkylnaphthalenes, alkylphenanthrenes, and triaromatic steroids. (K) to (N) are TICs of the same crude oil after the biodegradation experiments with different microorganism strains. (O) to (R) are TICs of the same crude oil after the biodegradation experiments with different bacteria strains and added surfactin (SF; surfactin). See Table 1 for identification of the bacteria used in each experiment. AlkylN = alkylnaphthalenes, P = phenanthrene, MP = methylphenanthrene, DMP = dimethylphenanthrene, TMP = trimethylphenanthrene, TAS = triaromatic steroid.

that showed the order of susceptibility to be: 3,4-DMDBT > 2,8DMDBT > 4,6-DMDBT [58]. The ease with which 3,4-DMDBT was biodegraded was attributed to it having an unsubstituted benzene ring that is preferentially attacked by aromatic hydrocarbon-degrading bacteria, whereas some other isomers have a methyl group on each of the homocyclic rings [58]. This order is different to other results previous reported. For example, using aerobic degradation experiments, it was shown that 1,2DMDBT and 1,9-DMDBT are more resistant to biodegradation than the other isomers [49]. Similarly, Jiménez et al.[15] showed that 4,6-DMDBT, 3,6-DMDBT, 1,4-DMDBT, 1,6-DMDBT and 1,8-DMDBT are more resistant to biodegradation, whereas the 2,4-DMDBT, 2,6-DMDBT, 3,7-DMDBT, and 1,3-DMDBT isomers are significantly depleted. Cheng et al. [11] observed that 3,6-DMDBT, 2,6-DMDBTand 2,4-DMDBT were biodegraded faster than 1,2-DMDBT in severely biodegraded reservoir oils, and that 4,6-DMDBT (ββ-substituted) is not the most readily biodegraded isomers. The discrepancies in the reported biodegradation of dimethyldibenzothiophenes suggest that selective biodegradation of PAHs is not solely controlled by their thermodynamic stability, but may also be

attacked and that fission of the ring devoid of the methyl group occurred [56]. There is no obvious variation in the MDBT/MP ratio with biodegradation (Table 3), indicting the rate of biodegradation of the methyldibenzothiophenes is similar to that of the methylphenanthrenes under the different conditions. However, the DBT/P ratio is higher in the samples treated with surfactin, suggesting that phenanthrene is more susceptible to biodegradation than dibenzothiophene under these experimental conditions, consistent with previous results [11,57]. The percentage depletion of the dimethyldibenzothiophene isomers ranges from 9.0 to 77.8% (Fig. 11), with a 1.5–54.3% increase in experiments with added surfactin. The biodegradability of the dimethyldibenzothiophenes is greater than that of the dimethylphenanthrenes (compare Figs. 8 and 11), so the dimethyldibenzothiophenes may be useful in assessing the extent of biodegradation in petroleum. The susceptibility of the dimethyldibenzothiophene isomers to microbial attack (from most susceptible to least; only for the non-co-eluting isomers) is: 1,2-DMDBT > 1,7DMDBT > 2,8-DMDBT > 3,6-DMDBT = 4,6-DMDBT > 2,6-DMDBT = 2,4-DMDBT. Similar results were also observed in laboratory experiments 9

Fuel 267 (2020) 117272

X. Wang, et al.

Fig. 7. Percentage loss due to biodegradation of phenanthrene (P) and the methylphenanthrenes (MP) in the crude oil samples after biodegradation, relative to the control. See Table 1 for identification of the bacteria used in each experiment.

related to their molecular configuration, as different organisms in mixed cultures probably follow different degradation pathways [49]. Additionally, PAH biodegradation rates are correlated with the relevant enzymes [59]. For example, Citrobacter sp strain J1 shows the highest degradation rate of dimethyldibenzothiophenes, and Brucella melitensis strain J2 prefers methyldibenzothiophenes. Other heteroatom compounds, such as carbazoles, benzocarbazoles and dibenzocarbazaoles, could be effectively used by the strain J2 with surfactin, as shown in previous experiments [21].

Distributions of C20, C21, C26, C27, and C28 triaromatic steroids identified in the aromatic fractions of the crude oils show that the C26, C27 and C28 isomers are all depleted by biodegradation at low rates relative to the control, varying from 2.6 to 32.2% for the studied bacteria (Fig. 12). The biodegradation rate was increased by addition of surfactin to 19.4–44.4%. The degradation rates of the short chain C21 and C20 triaromatic steroids are similar to the C26–C28 longchained counterparts. The C20-C21 triaromatic steroids have been shown to be preferentially removed during severe biodegradation Fig. 8. Percentage loss due to biodegradation of dimethylphenanthrenes (DMP) in the crude oil control and crude oil samples after biodegradation, relative to the control. 2,6-DMP+ = 2,6+ 2,7- + 3,5-dimethylphenanthrenes; 2,10DMP+ = 2,10- + 1,3- + 3,10- + 3,9-dimethylphenanthrenes; 1,6-DMP+ = 1,6- + 2,9+ 2,5-dimethylphenanthrenes; 4,9-DMP+ = 4,9- + 4,10- + 1,9-dimethylphenanthrenes. See Table 1 for identification of the bacteria used in each experiment.

10

Fuel 267 (2020) 117272

X. Wang, et al.

Fig. 9. Percentage loss due to biodegradation of trimethylphenanthrenes (TMP) in the crude oil control and crude oil samples after biodegradation, relative to the control. 1,3,6-TMP+ = 1,3,6- + 1,3,10- + 2,6,10-trimethylphenanthrenes; 1,3,7-TMP+ = 1,3,7- + 2,6,9- + 2,7,9trimethylphenanthrenes; 1,3,9-TMP+ = 1,3,9+ 2,3,6-trimethylphenanthrenes; 1,6,9-TMP+ = 1,6,9- + 1,7,9- + 2,3,7-trimethylphenanthrenes; 1,2,7-TMP+ = 1,2,7- + 1,2,9-trimethylphenanthrenes. See Table 1 for identification of the bacteria used in each experiment.

[8,11], but this does not seem to have occurred during this experiment. The methyltriaromatic steroids identified in the aromatic fractions of the crude oils have similar biodegradation rates to that of the triaromatic steroids (Fig. 13). The degradation rate of the methyltriaromatic steroids ranges from 6.1 to 28.9%, and surfactin addition increased the degradation rate by 6.3–31.5%. The high molecular weight methyltriaromatic steroids have been shown to be more resistant to biodegradation [8]. The Achromobacter strain J3 can degrade

the triaromatic steroids and methyltriaromatic steroids more than the other bacteria (Figs. 12 and 13). The biodegradation of the alkylphenanthrenes, alkyldibenzothiophenes, and methyltriaromatic steroids is clear in this study, but the biodegradation products of the alkyl substituents in crude oil need to be further studied. Many previous studies have shown the biodegradation products of individual methylphenanthrenes and alkyldibenzothiophenes, for example, the biodegradation products of 1-MP were salicylic acid and 6-methyl salicylic acid [60]. The major product of 2-MP

Fig. 10. Percentage loss due to biodegradation of dibenzothiophene (DBT) and methyldibenzothiophenes (MDBT) in the crude oil control and crude oil samples after biodegradation, relative to the control. See Table 1 for identification of the bacteria used in each experiment. 11

Fuel 267 (2020) 117272

X. Wang, et al.

Fig. 11. Percentage loss due to biodegradation of dimethyldibenzothiophenes (DMDBT) in the crude oil control and crude oil samples after biodegradation, relative to the control. 2,7DMDBT+ = 2,7- + 3,7-dimethyldibenzothiophenes; 1,4-DMDBT+ = 1,4- + 1,6- + 1,8-dimethyldibenzothiophenes; 1,3-DMDBT+ = 1,3+ 3,4-dimethyldibenzothiophenes; 2,3DMDBT+ = 2,3- + 1,9-dimethyldibenzothiophenes. See Table 1 for identification of the bacteria used in each experiment.

biodegradation was identified as the 2-methylphenanthrene aldehyde [61] or 1-hydroxy-6-methyl-2-naphtoic acid [62]. Oxidation products of methyldibenzothiophene isomers were methyl-HFBT and methylbenzothiophene-2,3-dione, dibenzothiophene methanols and some methyldibenzothiophene sulfones [56]. The metabolites of 3,4-DMDBT could be 6,7-dimethyl-3-hydroxy-2-formylbenzothiophene (HFBT), 6,7-dimethylbenzothiophene-2,3-dione, 6,7-dimethylbenzothiophene-2(3H)-one, or

3-hydroxy-6,7-dimethylbenzothiophene-2(3H)-one, varying by different bacteria, and the metabolite of 2,8-DMDBT was 5-methylbenzothiophene-2,3-dione [58]. In our previous resins fraction study, lower carbon numbers of the O2 class were detected in samples J3 and BC after biodegradation [21], which may be proof of the existence of these metabolites. Surfactin showed good enhancement in the biodegradation process of the above petroleum compounds, which proves that surfactin has

Fig. 12. Percentage loss due to biodegradation of triaromatic steroids (TAS) in the crude oil control and crude oil samples after biodegradation, relative to the control. 20R-C26 TAS + = C26 20R + C27 20S triaromatic steroids. See Table 1 for identification of the bacteria used in each experiment. 12

Fuel 267 (2020) 117272

X. Wang, et al.

Fig. 13. Percentage loss due to biodegradation of methyltriaromatic steroids (MTAS) in the crude oil control and crude oil samples after biodegradation, relative to the control. C27 3MTAS = C27 3-methyl triaromatic steroids, C27 4-MTAS = C27 4-methyl triaromatic steroid; C28 3,24-MTAS = C28 3-methyl-24 methyl triaromatic steroid; C29 4,23,24-MTAS = 4,23,24-trimethyl triaromatic steroid; C29 4 M,24EMTAS = C29 4-methyl-24-ethyl triaromatic steroid. See Table 1 for identification of the bacteria used in each experiment.

advantages in the removal of methylated aromatic compounds as well as of saturated hydrocarbons. This expands the scope of the application of surfactin in oil pollution remediation.

biodegradation of in-reservoir crude oils, mainly due to environmental conditions and bacteria types. 3.4. Determining the levels of biodegradation

3.3. Assessment of the order of susceptibility to microbial attack

Individual hydrocarbons within a compound class can be removed in a quasi-sequential order, and therefore, molecular parameters are also used to assess the level of biodegradation. According to the biodegradation scale proposed by Volkman et al. [4] and later modified by Peters and Moldowan [66] (abbreviated as the PM level), the levels of biodegradation caused by the different bacteria are very variable in the studied samples (Table 5). Sample BC is severely depleted in n-alkanes and n-alkylcyclohexanes, has an altered isoprenoid fraction (Figs. 1 and 3), and experienced partial loss of aromatic steroids (Figs. 12 and 13) and methylphenanthrenes (Fig. 7), suggesting biodegradation to PM level 3–4, without reaching full level 4 (alkylcyclohexanes removed). The Bacillus amyloliquefaciens strain BC is the best degrader of petroleum hydrocarbons, mainly because Bacillus has the ability to produce surfactin [23,24,26,28,31,32]. The other samples exposed to Citrobacter sp, Brucella melitensis and Achromobacter are slightly biodegraded, showing a general depletion of low molecular weight n-alkanes, and are classified as having reached PM level 1. Samples BC + SF and J3 + SF show a higher biodegradation level (n-alkanes and n-alkylcyclohexanes were removed, and isoprenoids reduced; (Figs. 1 and 3) than samples BC and J3, and are at PM level 4–5 [9]. Therefore, surfactin enhanced

The relative susceptibility to microbial attack of each of the compounds analysed in this study was determined from their observed depletion, as summarized in Table 4. There is a trend of lesser susceptibility to biodegradation with more substituents for the alkylphenanthrenes and alkyldibenzothiophenes in this study. Previous work has shown that the rate of biodegradation decreases with increasing number of alkyl substituents [4,5,7,11,49,63]. There is also a general trend for the preferential biodegradation of isomers with β-substituents, such as the 2-MP, 3-MP, 2-MDBT and 3-MDBT. PAH biodegradation may start by the oxidation of the aromatic ring [64] or the alkyl side chains [65]. Methylated phenanthrenes and dibenzothiophenes have been shown to be biodegraded at roughly the same rate [63], as is also shown here (Figs. 7 and 10). It is likely that the susceptibility of the different PAHs to biodegradation depends not only on their stereochemistry and thermodynamic considerations [9,11], but also on the microbial strain, substrate, substituent position, overall oil composition, nutrient availability, redox conditions and the species composition of the microbial population [20,49]. There are some differences between this study and previous studies of the

Table 3 Ratios based on aromatic hydrocarbons for the crude oil control and the crude oil samples after biodegradation. See Table 1 for identification of the bacteria used in each experiment. Ratio

Control

J1

J1 + SF

J2

J2 + SF

J3

J3 + SF

BC

BC + SF

MPI1 (3-MP + 2-MP)/(9-MP + 1-MP) DBT/P MDBT/MP

0.79 0.86 0.06 0.09

0.80 0.89 0.06 0.07

0.94 0.63 0.37 0.07

0.81 0.90 0.06 0.06

0.88 0.86 0.07 0.07

0.81 0.90 0.06 0.07

0.28 0.19 0.34 0.08

0.78 0.87 0.06 0.07

0.41 0.28 0.23 0.08

MPI-1 = 1.5 × [2-MP + 3-MP]/[P + 1-MP + 9-MP]; DBT/P = Dibenzothiophene/phenanthrene; MDBT/MP = (4-MDBT + 3-MDBT + 2-MDBT + 1-MDBT)/(3MP + 2-MP + 9-MP + 1-MP).

13

Fuel 267 (2020) 117272

X. Wang, et al.

Table 4 A summary of the order of susceptibility to microbial attack of alkylaromatic hydrocarbons during biodegradation. Compounds

Most susceptible Least susceptible

MP

P > MP > DMP > TMP MP: 2- > 3- = 1- > 9DMP: 1,2- > 1,8- = (2,6- +2,7- +3,5-) > 2,3- > 1,7- > (1,6- + 2,9- + 2,5-) = (4,9- + 4,10- + 1,9-) = (2,10- + 1,3- + 3,10- + 3,9-) TMP: (1,2,7- + 1,2,9-) > 1,2,6- > 2,3,10- = 1,3,8- > 1,2,8- = (1,3,7- + 2,6,9- + 2,7,9-) ~ (1,3,9- + 2,3,6-) = (1,6,9- + 1,7,9- + 2,3,7-) > 1,6,7- = (1,3,6+ 1,3,10- + 2,6,10-) DBT > MDBT > DMDBT MDBT: 2- +3- > 1- > 4DMDBT: 1,2- > 1,7- > 2,8- > 3,6- = 4,6- > 2,6- = 2,4-

MDBT

Abbreviations: P = phenanthrene; MP = methylphenanthrene; DMP = dimethylphenanthrene; TMP = trimethylphenanthrene; DBT = dibenzothiophene; MDBT = methyldibenzothiophene; DMDBT = dimethyldibenzothiophene. Table 5 Changes of hydrocarbon distribution during the biodegradation of the crude oil. Biodegradation markers

Samples

n-Alkanes n-Alkylcyclohexanes Isoprenoids C14-C16bicyclic terpane Hopanes Steranes 25-Norhopanes Diasteranes C26-C29 triaromatic steroids Methylphenanthrenes Dimethylphenanthrenes Trimethylphenanthrenes Methyldibenzothiophenes Dimethyldibenzothiophenes

J1

J1 + SF

J2

J2 + SF

J3

J3 + SF

BC

BC + SF

A U U U U U N A A A A A A A

SD SD A U U A N A A SD A A A A

A U U U U U N A A A A A A A

SD SD SD SD U U N A A SD A A A A

A U U U U A N A A A A A A A

SD SD A A U SD N A A SD A A SD SD

SD SD A A U U N A A A A A A A

SD SD A A U A N A A SD A A SD A

Abbreviations:U: Unaltered; A: Altered; SD: Severely depleted; N: No detected.

biodegradation on the PM scale by about 1–2 levels, based on the related hydrocarbon compositions.

dimethyldibenzothiophenes decrease in abundance with the level of biodegradation of the samples in a similar way to the methylphenanthrenes. A consistent sequence is observed with increasing biodegradation level of the oil samples examined. Thus, methyldibenzothiophenes and dimethyldibenzothiophenes are considered to be an indicator of aerobic biodegradation. Jiménez et al. [15] reported that removal of C1-and C2-phenanthrenes and dibenzothiophenes are good indicators of early biodegradation. C1-DBTs are more susceptible to biodegradation than C2-DBTs [63]. Levels of biodegradation ranking for aromatic hydrocarbons in crude oils has been extended here to

3.5. Alkyldibenzothiophenes as indicators of the extent of the biodegradation in petroleum It is evident from the above data that significant changes occur in the isomeric distribution of the methyldibenzothiophenes and dimethyldibenzothiophenes with increasing biodegradation (Figs. 10 and 11; Table 6). In general, the methyldibenzothiophenes and

Table 6 Effect of biodegradation on crude oils, based on the hydrocarbon composition including alkyldibenzothiophenes. Level of biodegradation

0 (Non-degraded) 1 (Minor) 3 (Moderate) 3–4 (Moderate) 4–5 (Moderate) 5–6 (Strong) 7 (Very extensive)

Chemical composition Saturated hydrocarbons [4]

Aromatic hydrocarbons, including alkylphenanthrenes [7,16]

Dibenzothiophene, methyldibenzothiophenes and dimethyldibenzothiophenes (this work)

n-Alkanes present General depletion of lower homologues of n-alkanes General n-alkanes depletion (> 90%) > 90% n-alkanes removed Isoprenoids slightly reduced Alkylcyclohexanes removed Isoprenoids reduced Bicyclic alkanes severely depleted Steranes affected

Unaltered Unaltered

Unaltered DBT, MDBT altered

MPs altered DMPs altered

DBT, MDBT altered DBT severely depleted MDBT altered MDBT severely depleted DMDBT altered MDBT absent DMDBT severely depleted

> 50% 5α,14α(H),17α(H)ethylcholestane-20R removed

MPs altered DMPs altered MPs absent DMPs altered EPs unaltered DMPs absent EPs absent

DMDBT severely depleted

Abbreviations: MP = methylphenanthrene; DMP = dimethylphenanthrene; EP = ethylphenanthrene; DBT = dibenzothiophene; MDBT = methyldibenzothiophene; DMDBT = dimethyldibenzothiophene. 14

Fuel 267 (2020) 117272

X. Wang, et al.

include alteration of dibenzothiophene, methyldibenzothiophenes and dimethyldibenzothiophenes (Table 6). The most biodegraded crude oil was completely depleted of methyldibenzothiophenes and dimethyldibenzothiophenes, suggesting that these components may be useful indicators at PM levels 5–6. However, some dimethyldibenzothiophenes are still present above PM levels 5–6.

Acknowledgements We thank the National Natural Science Foundation of China (Grant Nos. 41403067 and 41373126) and the State Scholarship Fund of the China Scholarship Council (File No. 201806445024) for their support. We would like to thank Dr. Shengbao Shi for technical support with the GC-MS. This paper was written whilst Xinwei Wang was a visiting Associate Professor at the Department of Earth and Planetary Sciences at Macquarie University, the support of which is gratefully acknowledged.

4. Conclusions This study highlights the effect of surfactin on the removal of methylaromatic hydrocarbons and saturated hydrocarbons during aerobic biodegradation experiments. Selective degradation of different hydrocarbons occurs in different ways for four types of bacteria. Bacillus amyloliquefaciens strain BC is the best for degrading petroleum hydrocarbons, mainly because it can produce surfactin. Achromobacter strain J3 can degrade methylphenanthrenes, methyltriaromatic steroids, triaromatic steroids and steranes better than the others. Citrobacter sp strain J1 causes the highest biodegradation rate of dimethylphenanthrenes, trimethylphenanthrenes, and dimethyldibenzothiophenes. Brucella melitensis strain J2 is good at degrading methyldibenzothiophenes. The removal rate of alkylaromatic and saturated hydrocarbons was significantly enhanced when surfactin was added to the samples in addition to the bacteria. Molecular parameters show that removal of aromatic hydrocarbons occurs during biodegradation of the crude oil, in addition to alteration of other petroleum hydrocarbons. A general loss of alkylation was observed in all samples, and the oxidation products of the dealkylation may be found in the resin fractions of crude oils. Alkyldibenzothiophenes and triaromatic steroids are useful indicators of the extent of biodegradation of petroleum. The order of susceptibility of alkylphenanthrenes, alkyldibenzothiophenes and methyltriaromatic steroids to biodegradation has been established, which is mainly consistent with previous observations from aerobic biodegradation experiments. These data partly contradict some reports of the biodegradation of in-reservoir crude oils, which is likely due to different bacteria and the dominantly anaerobic conditions subsurface. A consistent sequence of alteration of dibenzothiophene, methyldibenzothiophenes and dimethyldibenzothiophenes is observed with increasing level of biodegradation in the oil samples examined, thus these compounds are useful for determining the levels of biodegradation. The studied bacteria enabled light to moderate biodegradation of the crude oil, ranging to PM levels 3 to 4, whereas surfactin addition could enhance the amount of biodegradation by PM levels 1–2. Although many studies have focused on oil biodegradation and the effect of surfactin, this is one of the few studies that has described the effect of the biodegradation processes on alkylated aromatic compounds. Since surfactin increases the susceptibility of oil components to biodegradation, there is the potential to widely use surfactin under different environmental conditions during bioremediation.

Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.fuel.2020.117272. References [1] Howell VJ, Connan J, Aldridge AK. Tentative identification of demethylated tricyclic terpanes in nonbiodegraded and slightly biodegraded crude oils from the Los Llanos Basin, Colombia. Organic Geochem 1984;6:83–92. [2] Head IM, Jones DM, Larter SR. Biological activity in the deep subsurface and the origin of heavy oil. Nature 2003;426(6964):344–52. [3] Larter S, Wilhelms A, Head I, Koopmans M, Aplin A, Primio RD, et al. The controls on the composition of biodegraded oils in the deep subsurface—part 1: biodegradation rates in petroleum reservoirs. Org Geochem 2003;34(4):601–13. [4] Volkman JK, Alexander R, Kagi RI, Rowland SJ, Sheppard PN. Biodegradation of aromatic hydrocarbons in crude oils from the Barrow Sub-basin of Western Australia. Org Geochem 1984;6:619–32. [5] George SC, Boreham CJ, Minifie SA, Teerman SC. The effect of minor to moderate biodegradation on C5 to C9 hydrocarbons in crude oils. Org Geochem 2002;33(12):1293–317. [6] Dutta TK, Harayama S. Analysis of long-side-chain alkylaromatics in crude oil for evaluation of their fate in the Environment. Environ Sci Technol 2001;35(1):102–7. [7] Fisher S, Alexander R, Kagi R, Oliver G. Aromatic hydrocarbons as indicators of biodegradation in north Western Australian reservoirs. The Sedimentary Basins of Western Australia, vol. 2, Proceedings of Petroleum Exploration Society Symposium, Perth 1998:185–94. [8] Wardroper AMK, Hoffmann CF, Maxwell JR, Barwise AJG, Goodwin NS, Park PJD. Crude oil biodegradation under simulated and natural conditions—II. Aromatic steroid hydrocarbons. Organic Geochem 1984;6:605–17. [9] Huang H, Bowler BFJ, Oldenburg TBP, Larter SR. The effect of biodegradation on polycyclic aromatic hydrocarbons in reservoired oils from the Liaohe Basin, NE China. Organic Geochem 2004;35(11):1619–34. [10] Wang G, Xue Y, Wang D, Shi S, Grice K, Greenwood PF. Biodegradation and water washing within a series of petroleum reservoirs of the Panyu Oil Field. Org Geochem 2016;96:65–76. [11] Cheng X, Hou D, Mao R, Xu C. Severe biodegradation of polycyclic aromatic hydrocarbons in reservoired crude oils from the Miaoxi Depression. Bohai Bay Basin. Fuel 2018;211:859–67. [12] Cai M, Jiménez N, Krüger M, Guo H, Jun Y, Straaten N, et al. Potential for aerobic and methanogenic oil biodegradation in a water flooded oil field (Dagang oil field). Fuel 2015;141:143–53. [13] Díez S, Sabaté J, Viñas M, Bayona JM, Solanas AM, Albaigés J. The Prestige oil spill. I. Biodegradation of a heavy fuel oil under simulated conditions. Environ Toxicol Chem 2005;24(9):2203–17. [14] Franco MA, Viñas L, Soriano JA, de Armas D, González JJ, Beiras R, et al. Spatial distribution and ecotoxicity of petroleum hydrocarbons in sediments from the Galicia continental shelf (NW Spain) after the Prestige oil spill. Mar Pollut Bull 2006;53(5–7):260–71. [15] Jiménez N, Viñas M, Sabaté J, Díez S, Bayona JM, Solanas AM, et al. The Prestige oil spill. 2. Enhanced biodegradation of a heavy fuel oil under field conditions by the use of an oleophilic fertilizer. Environ Sci Technol 2006;40(8):2578–85. [16] Peters KE, Walters CC, Moldowan J. The biomarker guide. Cambridge University Press; 2005. [17] Douglas GS, Hardenstine JH, Liu B, Uhler AD. Laboratory and field verification of a method to estimate the extent of petroleum biodegradation in soil. Environ Sci Technol 2012;46(15):8279–87. [18] Ji G, Tan Y, Zhang L. Bacterial and granular sludge characteristics in an ultrahightemperature upflow anaerobic sludge blanket reactor treating super heavy oilcontaining wastewater. Environ Eng Sci 2010;28(2):129–37. [19] Greenwood PF, Wibrow S, George SJ, Tibbett M. Sequential hydrocarbon biodegradation in a soil from arid coastal Australia, treated with oil under laboratory controlled conditions. Org Geochem 2008;39(9):1336–46. [20] Lamberts RF, Christensen JH, Mayer P, Andersen O, Johnsen AR. Isomer-Specific Biodegradation of Methylphenanthrenes by Soil Bacteria. Environ Sci Technol 2008;42(13):4790–6. [21] Wang X, Cai T, Wen W, Zhang Z. Effect of biosurfactant on biodegradation of heteroatom compounds in heavy oil. Fuel 2018;230:418–29. [22] Xia W, Du Z, Cui Q, Dong H, Wang F, He P, et al. Biosurfactant produced by novel

CRediT authorship contribution statement Xinwei Wang: Conceptualization, Methodology, Resources, Funding acquisition, Writing - original draft. Ting Cai: Methodology, Validation, Writing - review & editing. Weitao Wen: Methodology, Visualization, Validation. Jiazhen Ai: Methodology, Validation. Jiayi Ai: Visualization. Zhihuan Zhang: Resources, Supervision, Funding acquisition. Lei Zhu: Methodology, Resources. Simon C. George: Supervision, Writing - review & editing. Declaration of Competing Interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. 15

Fuel 267 (2020) 117272

X. Wang, et al.

[23] [24]

[25] [26] [27] [28]

[29] [30] [31] [32] [33] [34] [35] [36] [37] [38] [39] [40] [41] [42]

[43] [44]

Pseudomonas sp. WJ6 with biodegradation of n-alkanes and polycyclic aromatic hydrocarbons. J Hazard Mater 2014;276:489–98. Bezza FA, Chirwa EMN. The role of lipopeptide biosurfactant on microbial remediation of aged polycyclic aromatic hydrocarbons (PAHs)-contaminated soil. Chem Eng J 2017;309:563–76. Nogueira Felix AK, Martins JJL, Lima Almeida JG, Giro MEA, Cavalcante KF, Maciel Melo VM, et al. Purification and characterization of a biosurfactant produced by Bacillus subtilis in cashew apple juice and its application in the remediation of oilcontaminated soil. Colloids Surf, B 2019;175:256–63. Banat IM, Franzetti A, Gandolfi I, Bestetti G, Martinotti MG, Fracchia L, et al. Microbial biosurfactants production, applications and future potential. Appl Microbiol Biotechnol 2010;87(2):427–44. Liu B, Liu J, Ju M, Li X, Yu Q. Purification and characterization of biosurfactant produced by Bacillus licheniformis Y-1 and its application in remediation of petroleum contaminated soil. Mar Pollut Bull 2016;107(1):46–51. Zhong H, Zeng GM, Yuan XZ, Fu HY, Haung GH, Ren FY. Adsorption of dirhamnolipid on four microorganisms. Appl Microbiol Biotechnol 2007;77(2):447–55. Hentati D, Chebbi A, Hadrich F, Frikha I, Rabanal F, Sayadi S, et al. Production, characterization and biotechnological potential of lipopeptide biosurfactants from a novel marine Bacillus stratosphericus strain FLU5. Ecotoxicol Environ Saf 2019;167:441–9. Liu Q, Lin J, Wang W, Huang H, Li S. Production of surfactin isoforms by Bacillus subtilis BS-37 and its applicability to enhanced oil recovery under laboratory conditions. Biochem Eng J 2015;93:31–7. Karlapudi AP, Venkateswarulu TC, Tammineedi J, Kanumuri L, Ravuru BK, Vr Dirisala, et al. Role of biosurfactants in bioremediation of oil pollution-a review. Petroleum 2018;4(3):241–9. Ali Khan AH, Tanveer S, Alia S, Anees M, Sultan A, Iqbal M, et al. Role of nutrients in bacterial biosurfactant production and effect of biosurfactant production on petroleum hydrocarbon biodegradation. Ecol Eng 2017;104:158–64. Long X, He N, He Y, Jiang J, Wu T. Biosurfactant surfactin with pH-regulated emulsification activity for efficient oil separation when used as emulsifier. Bioresour Technol 2017;241:200–6. Bordoloi NK, Konwar BK. Bacterial biosurfactant in enhancing solubility and metabolism of petroleum hydrocarbons. J Hazard Mater 2009;170(1):495–505. Jha SS, Joshi SJ. Lipopeptide production by Bacillus subtilis R1 and its possible applications. Braz J Microbiol 2016;47:955–64. Lai C-C, Huang Y-C, Wei Y-H, Chang J-S. Biosurfactant-enhanced removal of total petroleum hydrocarbons from contaminated soil. J Hazard Mater 2009;167(1):609–14. Darvishi P, Ayatollahi S, Mowla D, Niazi A. Biosurfactant production under extreme environmental conditions by an efficient microbial consortium, ERCPPI-2. Colloids Surf, B 2011;84(2):292–300. Cai T, Wang X, Wen W, Zhang Z, Wang T. Promotion effect of biosurfactant on heavy-oil biodegradation. Petrol Sci Technol 2017;35(16):1692–8. Volkman JK, Alexander R, Kagi RI, Woodhouse GW. Demethylated hopanes in crude oils and their applications in petroleum geochemistry. Geochim Cosmochim Acta 1983;47(4):785–94. Bence AE, Kvenvolden KA, Kennicutt MC. Organic geochemistry applied to environmental assessments of Prince William Sound, Alaska, after the Exxon Valdez oil spill—a review. Org Geochem 1996;24(1):7–42. Whyte LG, Bourbonnière L, Greer CW. Biodegradation of petroleum hydrocarbons by psychrotrophic Pseudomonas strains possessing both alkane (alk) and naphthalene (nah) catabolic pathways. Appl Environ Microbiol 1997;63(9):3719–23. Dzou LIP, Noble RA, Senftle JT. Maturation effects on absolute biomarker concentration in a suite of coals and associated vitrinite concentrates. Org Geochem 1995;23(7):681–97. Permanyer A, Márquez G, Gallego JR. Compositional variability in oils and formation waters from the Ayoluengo and Hontomín fields (Burgos, Spain). Implications for assessing biodegradation and reservoir compartmentalization. Org Geochem 2013;54:125–39. Hao R, Lu A. Biodegradation of heavy oils by halophilic bacterium. Prog Nat Sci 2009;19(8):997–1001. Snape I, Ferguson SH, Harvey PM, Riddle MJ. Investigation of evaporation and

[45] [46] [47]

[48] [49] [50] [51] [52] [53] [54] [55] [56] [57] [58] [59] [60] [61] [62] [63] [64] [65] [66]

16

biodegradation of fuel spills in Antarctica: II—Extent of natural attenuation at Casey Station. Chemosphere 2006;63(1):89–98. Bost FD, Frontera-Suau R, McDonald TJ, Peters KE, Morris PJ. Aerobic biodegradation of hopanes and norhopanes in Venezuelan crude oils. Org Geochem 2001;32(1):105–14. Shen H-H, Thomas RK, Penfold J, Fragneto G. Destruction and solubilization of supported phospholipid bilayers on silica by the biosurfactant surfactin. Langmuir 2010;26(10):7334–42. Bayona JM, Albaigés J, Solanas AM, Pares R, Garrigues P, Ewald M. Selective aerobic degradation of methyl-substituted polycyclic aromatic hydrocarbons in petroleum by pure microbial culturest. Int J Environ Anal Chem 1986;23(4):289–303. Rowland SJ, Alexander R, Kagi RI, Jones DM. Microbial degradation of aromatic components of crude oils: a comparison of laboratory and field observations. Org Geochem 1986;9(4):153–61. Budzinski H, Raymond N, Nadalig T, Gilewicz M, Garrigues P, Bertrand JC, et al. Aerobic biodegradation of alkylated aromatic hydrocarbons by a bacterial community. Org Geochem 1998;28(5):337–48. Cassani F, Eglinton G. Organic geochemistry of Venezuelan extra-heavy crude oils 2. Molecular assessment of biodegradation. Chem Geol 1991;91(4):315–33. Budzinski H, Nadalig T, Raymond N, Matuzahroh N, Gilewicz M. Evidence of two metabolic pathways for degradation of 2-methylphenanthrene by Sphingomonas sp. strain (2MPII). Environ Toxicol Chem 2000;19(11):2672–7. Nadalig T, Raymond N, Ni'matuzahroh, Gilewicz M, Budzinski H, Bertrand J. Degradation of phenanthrene, methylphenanthrenes and dibenzothiophene by a Sphingomonas strain 2mpII. Appl Microbiol Biotechnol 2002;59(1):79–85. Mazeas L, Budzinski H, Raymond N. Absence of stable carbon isotope fractionation of saturated and polycyclic aromatic hydrocarbons during aerobic bacterial biodegradation. Org Geochem 2002;33(11):1259–72. Permanyer A, Gallego JLR, Caja MA, Dessort D. Crude oil biodegradation and environmental factors at the riutort oil shale mine, Se Pyrenees. J Petroleum Geol 2010;33(2):123–39. Galarraga F, Urbani F, Escobar M, Márquez G, Martínez M, Tocco R. Main factors controlling the compositional variability of seepage oils from Trujillo State, Western Venezuela. J Pet Geol 2010;33(3):255–67. Saftic S, Fedorak PM, Andersson JT. Transformations of methyldibenzothiophenes by three Pseudomonas isolates. Environ Sci Technol 1993;27(12):2577–84. Ahmed M, Smith JW, George SC. Effects of biodegradation on Australian Permian coals. Org Geochem 1999;30(10):1311–22. Kropp KG, Andersson JT, Fedorak PM. Biotransformations of three dimethyldibenzothiophenes by pure and mixed bacterial cultures. Environ Sci Technol 1997;31(5):1547–54. Wammer KH, Peters CA. Polycyclic sromatic hydrocarbon biodegradation rates: a dtructure-based study. Environ Sci Technol 2005;39(8):2571–8. Zhong J, Luo L, Chen B, Sha S, Qing Q, Tam NFY, et al. Degradation pathways of 1methylphenanthrene in bacterial Sphingobium sp. MP9-4 isolated from petroleumcontaminated soil. Mar Pollut Bull 2017;114(2):926–33. Ni'matuzahroh, Gilewicz M, Guiliano M, Bertrand JC. In-vitro study of interaction between photooxidation and biodegradation of 2-methylphenanthrene by Sphingomonas sp. 2MPII. Chemosphere 1999;38(11):2501–7. Sabaté J, Grifoll M, Viñas M, Solanas AM. Isolation and characterization of a 2methylphenanthrene utilizing bacterium: identification of ring cleavage metabolites. Appl Microbiol Biotechnol 1999;52(5):704–12. Fedorak PM, Westlake DWS. Selective degradation of biphenyl and methylbiphenyls in crude oil by two strains of marine bacteria. Can J Microbiol 1983;29(5):497–503. Cerniglia CE. Biodegradation of polycyclic aromatic hydrocarbons. Curr Opin Biotechnol 1993;4(3):331–8. Mahajan MC, Phale PS, Vaidyanathan CS. Evidence for the involvement of multiple pathways in the biodegradation of 1- and 2-methylnaphthalene by Pseudomonas putida CSV86. Arch Microbiol 1994;161(5):425–33. Peters KE, Moldowan JM. The biomarker guide: Interpreting molecular fossils in petroleum and ancient sediments. United States: Englewood Cliffs, NJ (United States); Prentice Hall; 1993.