Cyclohexanecarboxylic acid degradation with simultaneous nitrate removal by Marinobacter sp. SJ18

Naphthenic acid (NA) is a toxic pollutant with potential threat to human health. However, NA transformations in marine environments are still unclear. In this study, the characteristics and pathways of cyclohexanecarboxylic acid (CHCA) biodegradation were explored in the presence of nitrate. The results showed that CHCA was completely degraded with pseudo-first-order kinetic reaction under aerobic and anaerobic conditions, accompanied by nitrate removal rates exceeding 70%, which was positively correlated with CHCA degradation (P < 0.05). In the proposed CHCA degradation pathways, cyclohexane is dehydrogenated to form cyclohexene, followed by ring-opening by dioxygenase to generate fatty acid under aerobic conditions or cleavage of cyclohexene through β-oxidation under anaerobic conditions. Whole genome analysis indicated that nitrate was removed via assimilation and dissimilation pathways under aerobic conditions and via denitrification pathway under anaerobic conditions. These results provide a basis for alleviating combined pollution of NA and nitrate in marine environments with frequent anthropogenic activities.


Introduction
Naphthenic acid (NA) is a ubiquitous functional group in the environment, mainly derived from crude oil, and can also be synthesized via the incomplete degradation of petroleum hydrocarbons by microorganisms or in situ degradation of petroleum (Brient et al. 2000;Whitby 2010). NAs and their metal salts are widely used in paint desiccants, corrosion inhibitors, emulsifiers, surfactants, lubricants, and wood preservatives, as well as to catalyze the production of alkyl and polyester resins (Clemente andFedorak 2005, Kannel andGan 2012). In recent years, NAs have been considered as toxic pollutants (Jie et al. 2015;Scarlett et al. 2013). NAs can penetrate the cell wall, destroy membrane lipid bilayers, or alter membrane properties due to their surfactant characteristics. It has been reported that NAs are acutely and chronically toxic to a variety of organisms, such as fish, amphibians, phytoplankton, and mammals (Frank et al. 2008;Kannel andGan 2012, Melvin andTrudeau 2012).
In general, NAs can migrate to terrestrial, aquatic, and marine environments via sewage discharge, crude oil leakage, precipitation runoff, and riverbank oil layer erosion (Kannel and Gan 2012;Scarlett et al. 2012). NAs have been Responsible Editor: Robert Duran Highlights • CHCA degradation in the presence of nitrate in marine environment was firstly studied • Strain SJ18 could rapidly degraded the combined pollutants of CHCA and nitrate • CHCA degradation pathways were proposed under aerobic and anaerobic conditions • Strain SJ18 contained functional genes for CHCA degradation and nitrate removal detected in aquatic systems, for example, the natural concentrations of NAs in the upper reaches of Athabasca River are generally < 1 mg L −1 (Schramm et al. 2000), while the NAs level can be reach as high as between 4 and 110 mg L −1 in the unconfined groundwater aquifers and tailings waters where oil sands are man-mined (Ahad et al. 2013;Hewitt et al. 2020;Kannel and Gan 2012). Furthermore, by employing improved method, NAs concentrations ranging from 2.29 to 132.91 mg kg −1 have been found in contaminated soil from oilfields in China (Wang et al. 2013(Wang et al. , 2015, and different types of NAs have been detected in seawater and sediments as a result of an oil tanker leakage in 2007 (Wan et al. 2014). In our previous study, 1-4 mg L −1 NAs were detected in the sediments of Dalian Bay, which was more than tenfold higher than the levels of polycyclic aromatic hydrocarbons (PAHs) (Zan et al. 2019).
To date, NAs attenuation in the environment includes physical adsorption, natural mineralization, plant degradation, and microbial degradation, of which microbial degradation is considered to be predominant (de Oliveira Livera et al. 2018, Huang 2011Lu et al. 2011;Quesnel et al. 2011). Numerous studies have explored the biodegradation of NAs by a variety of microorganisms, and the degradation pathways have long been ascertained (Blakley 1978;Del Rio et al. 2006a;Johnson et al. 2012;Presentato et al. 2018;Wang et al. 2015). For instance, Alcaligenes, Arthrobacter, and Corynebacterium have been found to degrade cyclohexanoic acid under aerobic conditions (Ougham and Trudgill 1982). The pathways proposed for the aerobic degradation of NAs mainly include β-oxidation, combined α-and β-oxidation, and aromatization pathways, with β-oxidation being the predominant biodegradation pathway (Quagraine et al. 2005;Whitby 2010). The α-and β-oxidation pathways have been detected during the aerobic transformation of cycloacetic acid by Alcaligenes sp. (Quagraine et al. 2005). Furthermore, analysis of hydronaphthoic acid degradation by sedimentary microorganisms revealed that the degradation process of decahydronaphthoic acid was completed by β-oxidation and Baeyer-Vil-liger oxidation (Del Rio et al. 2006b). The aromatization pathway was first described in Arthrobacter (PRL-W15), which could degrade cyclohexanoic acid by attacking alicyclic rings and form p-hydroxybenzoic acid through para-hydroxylation (Whitby 2010).
With regard to anaerobic biodegradation of NAs, the effects of different conditions (nitrate reduction, sulfate reduction, iron reduction, and methanogenesis) on anaerobic biodegradation of NAs have also been investigated (Clothier and Gieg 2016;Ghattas et al. 2017;Misiti et al. 2013). Clothier and Gieg (2016) found that some surrogate NAs (including cyclohexanecarboxylic acid (CHCA) and cyclohexaneacetic acid (CHAA)) were either fully or partially metabolized under nitrate-and sulfate-reducing conditions, while only CHAA was metabolized under methanogenic conditions and NAs were partially biodegraded in initial enrichments under ironreducing conditions. Some anaerobic microorganisms that could produce methane through β-oxidation were detected during the degradation of medium-and long-chain carboxylic acids; for example, microorganisms in the tailing sands environment could metabolize the cyclohexyl valerate side chain of 3-cyclohexyl propionic acid and 4-cyclohexyl butyric acid to methane (Holowenko et al. 2002). Gunawan et al. (2014) showed that a model surrogate NA could be readily metabolized under nitrate-reducing conditions in bioreactors.
However, it must be noted that these studies had predominantly focused on aerobic biodegradation of NAs in terrestrial environments, and knowledge about anaerobic degradation pathway and the potential for anaerobic NA biodegradation is still limited. Besides, biotransformation of NAs is even less understood in marine environment (Zan et al. 2022). It has long been recognized that nitrate pollution in coastal environment is mainly caused by intensive anthropogenic activities (Guo et al. 2020), and nitrate has been confirmed to be widely distributed in the offshore environment, coexisting with pollutants such as petroleum hydrocarbons and aromatic compounds (Capone and Bautista 1985, Hee-Sung Bae et al. 2002, Laufer et al. 2016. Studies have shown that the nitrate concentrations in the sediments of the Pearl River Estuary ranged from 6.6 to 92.1 mg kg −1 (Hong et al. 2019), while those in the groundwater of India were as high as 630.7 mg L −1 (Rina et al. 2013), indirectly posing a potential risk to the marine environment. In the Pearl River Delta, nitrate has been found to affect the biotransformation properties of microorganisms by interacting with a range of enzymes involved in metabolism or biodegradation (Su et al. 2018;Xu et al. 2014). Therefore, it is necessary to further study the biotransformation of NAs in marine environment, especially in the coastal areas that are highly affected by human activities.
In the present study, a facultative NA-degrading bacterium, widely distributed in the marine environment, was isolated from marine sediment, and its ability to transform surrogate NAs in the presence of nitrate under aerobic and anaerobic conditions was studied. Investigations of nitrate conversion, degradation intermediates, and draft genome sequencing were performed to explore NAs biodegradation characteristics in the marine environment as well as NAs biodegradation and nitrate removal pathways. The results obtained could help to improve our understanding of the characteristics of NAs biodegradation by marine bacteria.

Materials and methods
Surrogate NAs and microbial strain CHCA (C 7 H 12 O 2 , CAS: 98-89-5, purity: 99%, Meilunbio, China) was selected as a common model NA (Clothier and Gieg 2016;Demeter et al. 2014), and was dissolved in 0.1 mol L −1 NaOH solution to prepare a 10.0 g L −1 stock solution for later use. The microbial strain was obtained from the sediments of Dalian Bay (39°05′ N, 121°66′ E), China. The specific methods of strain enrichment and screening are provided in Supplementary Text S1.

Degradation of CHCA by the microbial strain
The CHCA degradation experiments were performed as follows: the microbial strain was pre-cultured in sterile 2216E medium for 24 h at 25 °C, and then subjected to centrifugation (10,000 × g, 10 min) in the logarithmic phase of growth (OD 600 = 1.5). The cell pellets were washed twice with sterile artificial sea water (ASW) and resuspended in 100-mL Erlenmeyer flasks containing 50 mL of ASW with 0.1 mL of trace medium (see Text S2 for ASW and trace medium composition) for aerobic CHCA degradation (the flasks were sealed with sterilized cotton stopper for oxygen exchange), and in 100-mL serum bottle for anaerobic degradation (the reaction system was exposed to highpurity He for 10-15 min to ensure an oxygen-free environment). The initial OD 600 of both the degradation systems was 0.05, initial CHCA concentration was 20 mg L −1 , and nitrate concentration was 50 mg L −1 . The experiments were performed in dark, and the pH was adjusted to 8.0 ± 0.2. The control groups were inactivated by high temperature (121 °C for 20 min), and all the experiments were performed in triplicate.

Extraction and detection of NAs
Samples were collected at 0, 12, 36, 60, 84, and 96 h of aerobic and anaerobic degradation of CHCA, respectively. Then, 1 mL of the collected samples was added to a 2-mL centrifuge tube containing 50 μL of NaOH (1 mol L −1 ), thoroughly mixed for 1 min, and centrifuged (8000 × g for 10 min). The supernatants obtained were collected and acidified with H 2 SO 4 (1 mol L −1 ) to pH < 2, and extracted thrice with dichloromethane (v/v = 1:1). The extracts were concentrated, dried with anhydrous Na 2 SO 4 , and quickly evaporated at 35 °C. Finally, the extracts were transferred to a 2-mL vial, and the solvent was evaporated in a gentle stream of nitrogen. Subsequently, 100 μL of derivatization reagent tert-butyldimethylsilyl (Sigma-Aldrich, St. Louis, MO, USA) were added to 100 μL of the extractive concentrates and the mixture was placed in a water bath at 60 °C for 20 min. Lastly, the derivatized samples were quantified by gas chromatography-mass spectrometry (GC-MS, Shimadzu Corporation GCMS-QP2020, Japan) (see Text S3 for detailed method).

Determination of chemical indices
Nitrate, nitrite, and ammonium contents were evaluated by sulfamic acid spectrophotometry, N-(1-naphthyl)-2-ethylenediamine spectrophotometry, and Nessler reagent spectrophotometry, respectively. The total organic carbon (TOC) was measured by an organic carbon meter (multifunctional N/C 3100, Analytik Jena GmbH, Germany), and the optical density of the microbial strain was determined by an ultraviolet spectrophotometer (SP-756P, Spectrum of Shanghai, China) at a wavelength of 600 nm (OD 600 ). The gaseous nitrogen was analyzed by gas chromatograph-thermal conductivity detector (GC7900-TCD, Techcomp of Shanghai, China). A total of 1 mL of headspace gas was injected, the temperature of the injection port and detector was 120 °C, and the carrier gas was high-purity He with a flow rate of 1.0 mL min −1 .

Strain identification and genome sequencing
The 16S rRNA of the microbial strain was amplified by the general PCR method of prokaryotic bacteria, and the gene primers 27F (5'-AGA GTT TGA TCC TGG CTC AG-3') and 1492R (5'-GGT TAC CTT GTT ACG ACT T-3') were used. The PCR conditions were as follows: pre-denaturation at 94 °C for 5 min, followed by 35 cycles of denaturation at 94 °C for 1 min, annealing at 55 °C for 1 min, and extension at 72 °C for 1 min, and a final repair and extension at 72 °C for 10 min. The 16S rRNA gene sequence was compared in NCBI (http:// www. ncbi. nlm. nih.), and the phylogeny was established using MEGA5 software program (https:// mega51. softw are. infor mer. com/).
The Illumina HiSeq 2500 sequencing platform and PE150-100X library-building sequencing strategy were employed for draft genome sequencing. The clean data were obtained by removing the adapters and processing lowquality sequences from the raw data. SOAP denovo version 2.04 software was employed to assemble the sequence, select the best assembly result, and fill the hole and optimize the assembly result (Lim et al. 2016), and all these processes were accomplished by Novogene Technology Co., Ltd. (see Text S4 for specific methods).

Characteristics of CHCA degradation and nitrate removal
The isolated microbial strain, labeled as strain SJ18, was grown in 2216E medium and then inoculated onto 2216E solid agar plate containing 50 mg L −1 NAs to examine its colony characteristics. The colony appeared round, translucent, and white on the 2216E solid agar plate (Fig. S1). The 16S rRNA sequencing results showed that the similarity between strain SJ18 and Marinobacter profundi was > 99%. The gene size of strain SJ18 was 4,364,528 bp, a total of 3951 genes were annotated, and the average GC content was 59.84%. Based on these characteristics, strain SJ18 was identified as Marinobacter sp. strain SJ18 (GenBank Accession No. MH458950.1). Figure S2 of Supplementary Material illustrates the phylogenetic tree of strain SJ18.
The aerobic and anaerobic degradation of CHCA are shown in Fig. 1a, b. Under aerobic conditions, CHCA was rapidly degraded after the start of the experiment, and was completely degraded within 60 h. The maximum degradation rate appeared at around 24 h (degradation rate of about 0.43 mg h −1 ). In contrast, no decrease in CHCA was noted in the control (heat-killed group), indicating that abiotic factors have negligible influence on aerobic CHCA degradation. Moreover, the TOC content in the experimental group decrease from 18.9 ± 0.5 to 7.0 ± 0.1 mg L −1 , and the OD 600 gradually increased with the progress of degradation, reaching a maximum of 0.74 at the end of CHCA degradation (about 60 h). Subsequently, the OD 600 gradually decreased, suggesting that strain SJ18 could utilize CHCA as the only carbon source and that the cell density slowly decreased after complete consumption of CHCA. In addition, the nitrate concentration gradually decreased, ammonium concentration increased, and nitrite concentration showed no significant changes in the experimental group. The nitrate removal rate was > 70% throughout the experiment, with the rate of nitrate consumption being consistent with that of aerobic degradation of CHCA, and no nitrogen was captured during the degradation process, indicating that nitrate was reduced to ammonium, which was utilized by the bacterial cells. In contrast, no significant changes in the corresponding indicators were noted in the blank control group.
Under anaerobic conditions, CHCA was completely degraded within 84 h, with maximum degradation rate of about 0.37 mg h −1 (around 36 h), implying that the anaerobic degradation efficiency was lower than aerobic degradation efficiency. The TOC concentration in the experimental group decreased from 18.7 ± 0.3 to 8.5 ± 0.1 mg L −1 , and the change in OD 600 value was consistent with that noted under aerobic conditions. However, nitrite concentration continued to increase with the consumption of nitrate, and no significant changes were detected in the ammonium content. At the end of the CHCA degradation process, the nitrite concentration gradually decreased and nitrogen production was detected (data not shown). The nitrate removal rate reached 70%, and the nitrate consumption rate was consistent with the CHCA anaerobic degradation rate. These findings suggested the occurrence of denitrification during anaerobic degradation of CHCA. The fitting results of degradation curves showed that aerobic and anaerobic degradation of CHCA conformed to the pseudo-first-order kinetic reaction (Fig. S3). The reaction rate constant k was 0.0342 (R 2 = 0.936) and 0.0292 (R 2 = 0.946) for aerobic and anaerobic degradation, respectively. Pearson correlation analysis showed that both aerobic and anaerobic degradation of CHCA were significantly positively correlated with nitrate removal, with correlation coefficients of 0.871 (P < 0.05) and 0.843 (P < 0.05), respectively.

Intermediates and pathways of CHCA degradation
To explore the aerobic and anaerobic degradation pathways of CHCA, samples collected at 0, 36, 60, and 84 h of the experiment were subjected to GC-MS to detect the intermediate products. The results showed several intermediate products (Fig. 2), and the mass spectral data revealed the presence of cyclohexenecarboxylic acid (m/z = 198, 183) in both aerobic and anaerobic groups. The appearance of cyclohexenecarboxylic acid indicated that CHCA might have lost two electrons through dehydrogenation and formed a carbon-carbon double bond on the cyclohexane ring. This degradation pathway is generally considered to be typical β-oxidation and most of the aromatic or cycloalkane pollutants are known to be degraded via this pathway Whitby 2010). Glycerol (m/z = 205,147,73) was detected in the aerobic group, while lactic acid (m/z = 191, 147, 117) was found in the anaerobic group. Glycerol and lactic acid are presumed to be the oxidative hydrolysis products of fatty acids under aerobic and anaerobic conditions, respectively, and their presence implied that hydrolysis could open the ring of cyclohexenecarboxylic acid and eventually form short-chain fatty acids. However, no specific fatty acid products were detected during the degradation process, which might be owing to the fact that fatty acids are readily utilized by microorganisms, resulting in their short life during degradation, and this speculation has also been confirmed in some reports. Quesnel et al. (2011) found that an intermediate product of CHCA existed for a short period of time during the cyclohexylacetic acid biodegradation process and could not be captured. Nevertheless, although the product after ring opening could not be captured, it could be essentially confirmed that CHCA was mainly biodegraded through β-oxidation.
The whole genome sequencing was also employed to explore the CHCA degradation process. The COG, GO, and KEGG databases of strain SJ18 were analyzed to predict the biological process, functional genes, and degradation pathways of CHCA under aerobic and anaerobic conditions. The COG analysis results (Fig. S4) showed that the functional genes, including those involved in energy production and conversion, amino acid transport and metabolism, carbohydrate transport and metabolism, coenzyme transport and metabolism, lipid transport and metabolism, and cell wall/membrane/envelope biogenesis, were significantly enriched. The results of GO analysis (Fig. S5) revealed that the enriched genes in the genome of strain SJ18 were mainly involved in cellular process, metabolic process, binding, catalytic activity, nitrogen utilization, antioxidant activity, enzyme regulator activity, transporter activity, and other functions. The KEGG annotation results (Fig. 3) demonstrated that the pathways of membrane transport, coenzyme metabolism, carbohydrate metabolism, and amino acid metabolism were also enriched. In addition, pathways such as ABC transporters, fatty acid degradation and biosynthesis, degradation of ketone bodies, pyruvate metabolism and styrene degradation, chlorocyclohexane and chlorobenzene degradation, benzoate degradation, toluene degradation, and naphthalene degradation were also obviously enriched in strain SJ18. Some genes associated with CHCA degradation in these pathways, such as genes encoding long-chain acyl-CoA synthase (fadD), benzoate 1,2-dioxygenase (benA), catechol 1,2-dioxygenase (catA), 3-hydroxyacyl-CoA dehydrogenase (HADH), hydroxycyclohexene carboxylic acid dehydrogenase (benD), alcohol dehydrogenase (adh), and 3-hydroxy-3-methylglutaryl coenzyme A lyase (hmgL), have been confirmed to participate in β-oxidation and metabolism of alkanes and PAHs under aerobic and anaerobic conditions Cameron et al. 2019;Kung et al. 2013;McKew et al. 2021). These results indicated that the abovementioned functional genes may play an important role in the biodegradation of CHCA. Subsequently, the aerobic and anaerobic degradation pathways of CHCA were inferred based on the abovementioned findings as follows (Fig. 4): First, two hydrogen atoms from CHCA were removed by dehydrogenase, forming a carbon-carbon double bond (cyclohexenecarboxylic acid) on the cyclohexane ring. In the aerobic degradation of CHCA, the cyclohexenecarboxylic acid ring was opened by specific dioxygenase (such as benzoate 1,2-dioxygenase or catechol 1,2-dioxygenase) in the presence of oxygen to form fatty acids. In the anaerobic degradation of CHCA, cyclohexenecarboxylic acid was hydrolyzed and dehydrogenated to sequentially generate 1-hydroxy-CHCA and cyclohexanonecarboxylic acid, which were subsequently hydrolyzed and opened to form short-chain fatty acids. Finally, glycerol and lactic acid were formed through multiple β-oxidation under aerobic and anaerobic conditions, respectively.

Nitrate metabolism pathway during CHCA degradation
Nitrate reduction to ammonia and denitrification were considered to accompany aerobic and anaerobic degradation of CHCA, respectively. Hence, the transformation of nitrate was further elucidated using genome sequencing. The genome sequence of strain SJ18 was compared with the KEGG database, and the nitrogen metabolism pathway of strain SJ18 was determined (Fig. 5). A total of seven genes related to nitrogen metabolism were found, including nitrate reductase (NarGHI), nitrate assimilation reduction catalytic subunit (NasAB), nitrite reductase (NirS), nitrite reductase subunits (NirBD), nitric oxide reductase (NorBC), nitrous oxide reductase (NosZ), and nitrite oxidoreductase (NxrAB). Among them, NarGHI, NirS, NorBC, and NosZ are considered as common denitrification genes that could perform complete denitrification process, whereas NarGHI and NirBD could convert nitrate to ammonium by dissimilatory reduction (Lu  Marchant et al. 2017). However, NarGHI is oxygen-sensitive and its expression is inhibited under aerobic conditions . As strain SJ18 lacks gene encoding NapAB (periplasmic nitrate reductase, which can reduce nitrate to nitrite under aerobic conditions), nitrate is converted to nitrite by NasAB under aerobic conditions.
Based on the inorganic nitrogen indicators and genomic results, the nitrate transformation process during aerobic and anaerobic biodegradation of CHCA was proposed as follows: under aerobic conditions, nitrate as an electron donor is transported from extracellular to intracellular region through nitrate transporter (Nrt), and NasAB reduces nitrate to nitrite. Then, nitrite is reduced to ammonium by NirBD. Finally, ammonium is transformed to glutamate by glutamine synthase (glnA) and glutamate dehydrogenase (gudB/rocG), and glutamate is subsequently metabolized. Under anaerobic conditions, nitrate is reduced to nitrite by NarGHI, nitrite is further reduced to nitric oxide (NO) by NirS, NO is reduced to nitrous oxide by NorBC, and nitrous oxide is reduced to nitrogen by NosZ.
In general, strain SJ18 not only exhibited the ability to reduce nitrate to ammonium via dissimilated nitrate reduction to ammonium (DNRA) pathway and assimilated nitrate reduction to ammonium (ANRA) pathway, but also showed denitrification ability. During the nitrate removal process, the first step of the DNRA pathway is denitrification by NarGHI to reduce nitrate. In the ANRA pathway, NasAB assimilates and reduces nitrate to nitrite. However, owing to the lack of NirA-or NIT-6-coding genes, strain SJ18 was unable to reduce nitrate to ammonium through the ANRA pathway. Therefore, the nitrite produced in the first step of the ANRA pathway had to be further metabolized through other nitrogen metabolism pathways such as denitrification or the DNRA pathway. In addition, strain SJ18 also encoded nitrite oxidoreductase, which could nitrify nitrite to form nitrate.
Overall, the nitrate metabolic pathways under aerobic and anaerobic biodegradation of CHCA by strain SJ18 were obviously different. Under aerobic conditions, nitrate was gradually consumed with the degradation of CHCA, along with cell growth (increase in OD value), indicating that nitrate was converted to ammonium via the ANRA and DNRA pathways, and that the ammonium produced was used for cell growth. Under anaerobic conditions, the nitrite concentration rapidly increased with nitrate consumption, and nitrogen was detected at the end of the process. Denitrification is considered to be an important pathway in nitrate metabolism under anaerobic conditions. It is believed that nitrite accumulation in the degradation system might be owing to the faster rate of nitrate reduction (by NarGHI) than nitrite reduction (by NirS), NO reduction (by NorBC), and nitrous oxide reduction (by NosZ).

Conclusion
This study demonstrated that CHCA was completely degraded by Marinobacter sp. SJ18 within 60 and 84 h under aerobic and anaerobic conditions, respectively, and a significant positive correlation was found between nitrate removal and CHCA degradation. The CHCA degradation pathways were speculated as follows: cyclohexane ring is first dehydrogenated to form cyclohexene, and the cyclohexene carboxylic acid ring is opened via the action of dioxygenase to generate fatty acid under aerobic conditions and via β-oxidation under anaerobic conditions. In addition, nitrate is utilized through ANRA and DNRA pathways under aerobic conditions and removed by denitrification under anaerobic conditions. Thus, this study provides a basis for alleviating combined pollution of NAs and nitrate in marine environment with frequent anthropogenic activities, although the behavior and biotransformation of NAs in marine environment require further exploration.
Author contribution Shuaijun Zan: investigation, experimental design, methodology, writing-original draft, writing-review/editing. Jing Wang: planned research, involved in supervised all analyses, data interpretation and discussion as a Project Leader. Jingfeng Fan: participated in project cooperation. Yuan Jin: assisted in conducting some experiments. Zelong Li: assisted in conducting some sample analyses and writing editing. Miaomiao Du: assisted in conducting some experiments.