Determination of soil phenanthrene degradation through a fungal–bacterial consortium

Chunling Luo; Guoqing Guan; Yeliang Dai; Xixi Cai; Qihui Huang; Jibing Li; Gan Zhang

doi:10.1128/aem.00662-24

. 2024 May 16;90(6):e00662-24. doi: 10.1128/aem.00662-24

Determination of soil phenanthrene degradation through a fungal–bacterial consortium

Chunling Luo ^1,², Guoqing Guan ^1,², Yeliang Dai ^1,², Xixi Cai ³, Qihui Huang ⁴, Jibing Li ^1,^2,^✉, Gan Zhang ^1,²

Editor: Irina S Druzhinina⁵

PMCID: PMC11218650 PMID: 38752833

ABSTRACT

Fungal–bacterial consortia enhance organic pollutant removal, but the underlying mechanisms are unclear. We used stable isotope probing (SIP) to explore the mechanism of bioaugmentation involved in polycyclic aromatic hydrocarbon (PAH) biodegradation in petroleum-contaminated soil by introducing the indigenous fungal strain Aspergillus sp. LJD-29 and the bacterial strain Pseudomonas XH-1. While each strain alone increased phenanthrene (PHE) degradation, the simultaneous addition of both strains showed no significant enhancement compared to treatment with XH-1 alone. Nonetheless, the assimilation effect of microorganisms on PHE was significantly enhanced. SIP revealed a role of XH-1 in PHE degradation, while the absence of LJD-29 in ¹³C-DNA indicated a supporting role. The correlations between fungal abundance, degradation efficiency, and soil extracellular enzyme activity indicated that LJD-29, while not directly involved in PHE assimilation, played a crucial role in the breakdown of PHE through extracellular enzymes, facilitating the assimilation of metabolites by bacteria. This observation was substantiated by the results of metabolite analysis. Furthermore, the combination of fungus and bacterium significantly influenced the diversity of PHE degraders. Taken together, this study highlighted the synergistic effects of fungi and bacteria in PAH degradation, revealed a new fungal–bacterial bioaugmentation mechanism and diversity of PAH-degrading microorganisms, and provided insights for in situ bioremediation of PAH-contaminated soil.

IMPORTANCE

This study was performed to explore the mechanism of bioaugmentation by a fungal–bacterial consortium for phenanthrene (PHE) degradation in petroleum-contaminated soil. Using the indigenous fungal strain Aspergillus sp. LJD-29 and bacterial strain Pseudomonas XH-1, we performed stable isotope probing (SIP) to trace active PHE-degrading microorganisms. While inoculation of either organism alone significantly enhanced PHE degradation, the simultaneous addition of both strains revealed complex interactions. The efficiency plateaued, highlighting the nuanced microbial interactions. SIP identified XH-1 as the primary contributor to in situ PHE degradation, in contrast to the limited role of LJD-29. Correlations between fungal abundance, degradation efficiency, and extracellular enzyme activity underscored the pivotal role of LJD-29 in enzymatically facilitating PHE breakdown and enriching bacterial assimilation. Metabolite analysis validated this synergy, unveiling distinct biodegradation mechanisms. Furthermore, this fungal–bacterial alliance significantly impacted PHE-degrading microorganism diversity. These findings advance our understanding of fungal–bacterial bioaugmentation and microorganism diversity in polycyclic aromatic hydrocarbon (PAH) degradation as well as providing insights for theoretical guidance in the in situ bioremediation of PAH-contaminated soil.

KEYWORDS: stable-isotope probing, fungal–bacterial bioaugmentation, extracellular enzyme activity, active-degrading microorganisms, PHE assimilation

INTRODUCTION

Polycyclic aromatic hydrocarbons (PAHs) are a class of persistent organic pollutants with carcinogenic, teratogenic, and mutagenic properties (1, 2). They are widespread in the environment and can be found in various aquatic and terrestrial ecosystems (3 –5). These compounds primarily originate from incomplete combustion of fossil fuels (coal and oil) and oil spills associated with crude oil production (6, 7). Considering the detrimental effects of such soil pollution on human health and the environment, efficient remediation measures are needed. Biological remediation of PAH-contaminated soil offers environmental and cost advantages over chemical and physical methods, making it a promising restoration technique (8). In native habitats, indigenous functional microorganisms play crucial roles in the natural attenuation of soil organic pollutants (9, 10). Bioaugmentation, a commonly used enhancement method in the bioremediation of PAHs, involves the introduction of a single strain or microbial consortium to enhance the biodegradation capacity of contaminated sites and has been successfully validated for soil PAH remediation (11). However, the application of this method faces significant challenges due to variations in soil physicochemical properties and competition from indigenous microorganisms (12, 13). Consequently, introducing autochthonous microorganisms has been proposed as a means to effectively address these issues and improve the efficacy of bioaugmentation (14, 15).

Numerous PAH-degrading bacteria and fungi have been identified and used for in situ bioaugmentation (16, 17). As major soil microorganisms, fungi exhibit robust survival capabilities and secrete enzymes with low-substrate specificity, enabling them to degrade various PAHs (18). Their long-distance transport abilities enhance their degradation of a wide range of pollutants (17). Fungi predominantly degrade PAHs using cytochrome P450 monooxygenases and an extracellular ligninolytic enzyme system, involving manganese peroxidases and lignin peroxidases to break down recalcitrant pollutants (19). In contrast, bacteria often degrade PAHs through PAH dioxygenases (20). However, these findings were mainly from laboratory-based pure culture studies, and in situ bioaugmentation of fungi and bacteria in real environments may involve unique and complex mechanisms. The biodegradation efficiency of autochthonous bioaugmentation is primarily attributed to changes in microbial community structure and the enrichment of PAH-degrading bacteria, as well as PAH degradation-related genes (13, 17, 21). However, the true collaborative mechanisms of action of indigenous fungi and bacteria in the in situ PAH biodegradation in contaminated soil are unclear, including whether added fungi and bacteria can enhance the degradation or assimilation of PAHs and their involvement in this process, as well as the impact of their combined effects on microbial diversity and functional bacterial and fungal communities. Therefore, investigating the effects of combined fungi and bacteria on PAH biodegradation and elucidating their roles and relations with indigenous functional microorganisms will provide comprehensive insights into the mechanisms of bioaugmentation.

Although various soil PAH-degrading microbes have been identified through pure culture methods, challenges remain in identifying active degraders in complex environments and confirming the functionality of added microorganisms (22 –24). Stable isotope probing (SIP) is a culture-independent technique involving the labeling of environmental microorganisms by the incorporation of heavy isotopes (¹³C, ¹⁵N, or ¹⁸O) into substrates (25, 26). This technique has successfully identified and traced active functional bacteria and has been used to investigate their diversity and changes in activity during bioaugmentation (13, 14). However, there have been no studies regarding the co-biodegradation mechanisms of action of fungi and bacteria for organic pollutants or tracking of their functionalities.

In this study, we isolated one highly efficient indigenous PAH-degrading fungus and bacterium from petroleum-contaminated soil. We hypothesized that there are distinct remediation mechanisms between fungi and bacteria in PAH-contaminated soil. Unlike bacteria, fungi may not directly assimilate PAHs but instead facilitate PAH transformation through the secretion of extracellular oxidases. Furthermore, fungi may facilitate bacterial assimilation of pollutants without accelerating their removal. Using phenanthrene (PHE) as a model PAH compound commonly found in the environment, this study was performed to investigate PAH biodegradation by autochthonous fungi and bacteria in petroleum-contaminated soil. We used DNA-SIP and high-throughput sequencing to elucidate their combined mechanisms. This study provided new insights into the bioremediation mechanisms of fungi and bacteria in soil contaminated with organic pollutants.

RESULTS

PHE degradation

The PHE degradation efficiencies increased over time for all biotic treatments (Fig. 1). However, the PHE content in sterilized soil treatment (ST) showed minimal changes, suggesting limited self-removal capability of PHE without microbial activity. Microorganisms play vital roles in the degradation of soil PHE. After 8 days of incubation, the control treatments including NC (negative control using natural soil) and NC-S (natural soil with fungal agent substrate) treatments showed PHE degradation efficiencies of 16.2% and 18.1%, respectively. Compared to the controls, bioaugmentation treatments with Pseudomonas XH-1 (Bac, 63.6%), Aspergillus LJD-9 (Fun, 33.3%), and both strains combined (Com, 70.3%) showed significantly enhanced PHE removal efficiency (P < 0.05). However, the degradation effect of Fun treatment was not as significant as that of Bac treatment, while Com treatment showed the highest PHE degradation efficiency. These observations suggested that the combined use of XH-1 and LJD-29 was more effective in PHE degradation compared to either stain alone, with Pseudomonas XH-1 having a greater impact.

PHE-degrading bacteria and fungi revealed by DNA-SIP

Prior to DNA ultracentrifugation, 16S rRNA and internal transcribed spacer (ITS) amplicon sequencing were conducted to analyze bacterial and fungal communities. A total of 16,296,156 bacterial amplicon sequence variants (ASVs) and 10,347,828 fungal ASVs were identified, and the composition was analyzed at the ASV level, focusing on the top 10 phyla (Fig. S1). In NC treatment, the dominant bacterial phyla (>5%) were Gammaproteobacteria (38.3%), Actinobacteria (30.4%), and Alphaproteobacteria (22.5%). However, the growth of Actinobacteria was inhibited, and the community composition changed in the NC-S, Fun, Bac, and Com treatments. Moreover, the Fun and Com treatments significantly increased the abundance of Alphaproteobacteria (62.4% and 49.9%, respectively). With regard to fungal communities, Eurotiomycetes (78.2%) and Sordariomycetes (17.9%) were the dominant phyla in the NC treatment. In the NC-S treatment, Eurotiomycetes (42.5%), Sordariomycetes (43.9%), and Mucoromycota (12.1%) were the main phyla. However, the growth of Mucoromycota was significantly inhibited, and the community composition changed in the Fun, Bac, and Com treatments. Furthermore, both Fun and Com treatments showed similar increases in the abundance of Eurotiomycetes (approximately 86.4%).

To distinguish the light and heavy DNA fractions in different SIP treatments, real-time PCR (qPCR) was performed to quantify 16S rRNA and ITS genes in fractions with different buoyant density (BD). Figure S2 shows significant enrichment of 16S rRNA genes in the heavy DNA fractions of the ¹³C-PHE treatments. Moreover, the bioaugmentation treatments (Bac, Fun, and Com) exhibited a notably higher relative enrichment factor (REF) value (5.61) compared to the non-bioaugmentation treatments (NC and NC-S), suggesting that bioaugmentation significantly enhanced the effectiveness of microbial assimilation. However, the ITS gene did not show significant enrichment in the heavy DNA fractions of all ¹³C-PHE microcosms. Following REF detection, a total of 13 functional PHE-degrading bacteria were screened and identified (Fig. 2; Table S2). However, no active functional fungi were detected, suggesting that they had no direct involvement in PHE assimilation.

Fig 2 — Heatmap of key ASVs in the heavy DNA fractions from the NC, NC-S, Fun, Bac, and Com treatments. NC, negative control using natural soil; NC-S, natural soil with fungal agent substrate; Bac, bioaugmentation with bacterial *Pseudomonas* XH-1; Fun, bioaugmentation with *Aspergillus* sp. LJD-29 agent; Com, bioaugmentation with both LJD-29 and XH-1; ST, sterilized soil. Colors indicate the REF (a) and relative abundance (b), ranging from purple (low) to red (high). Values are means of three independent replicates.

Figure 3 shows the phylogenetic information of active microorganisms, with identified functional ASVs from Actinobacteria, Alphaproteobacteria, and Gammaproteobacteria. During NC treatment, five enriched ASVs were identified, i.e., Mycobacterium (ASV_59 and ASV_77), Mycolicibacterium (ASV_107), Pseudomonas (ASV_4), and Ensifer (ASV_39). The NC-S treatment resulted in the detection of two PHE degraders from Marinobacter (ASV_17) and Alcanivorax (ASV_44). Bioaugmentation altered the composition of PHE-degrading microbial communities. In Bac treatment, Pseudomonas (ASV_4), Nocardioides (ASV_37), Mycobacterium (ASV_59 and ASV_107), Micromonospora (ASV_88), and Ramlibacter (ASV_89) were identified. Pseudomonas (ASV_4), Lysobacter (ASV_28), and Mycobacterium (ASV_77) were enriched as active degraders in the Fun microcosm. In the Com microcosm, Pseudomonas (ASV_4), Streptomyces (ASV_13 and ASV_42), and Micromonospora (ASV_88) were identified. Notably, ASV_4, ASV_59, ASV_77, and ASV_107 were common among treatments, suggesting that they have significant roles in PHE degradation. It is noteworthy that the functional microorganisms represented by ASV_4 showed 100% 16S rRNA gene sequence similarity to Pseudomonas XH-1. This confirmed that ASV_4 represented the added Pseudomonas XH-1, and their involvement in PHE degradation was observed in all treatments except NC-S. Moreover, the abundance of ASV_4 in the heavy DNA of the Bac and Com microcosms was significantly higher (20.9% and 5.20%, respectively) compared to the NC treatment (3.81%), highlighting the importance of strain XH-1 in the in situ degradation of PHE.

Fig 3 — Phylogenetic tree of ASVs responsible for *in situ* PHE degradation based on the neighbor-joining method using 16S rRNA gene sequences. The bar represents 0.05 substitutions per nucleotide position. The active PHE degraders identified from NC, NC-S, Fun, Bac, and Com treatments are highlighted in red, yellow, green, blue, and gray, respectively. NC, negative control using natural soil; NC-S, natural soil with fungal agent substrate; Bac, bioaugmentation with bacterial *Pseudomonas* XH-1; Fun, bioaugmentation with *Aspergillus* sp. LJD-29 agent; Com, bioaugmentation with both LJD-29 and XH-1; ST, sterilized soil.

The abundance of Aspergillus LJD-29 and Pseudomonas XH-1 and their effects on PHE degradation

As shown in Fig. 4, indigenous strains LJD-29 (65.6%) and XH-1 (5.21%) dominated the natural soil (NC; negative control). The NC-S treatment increased the abundance of strain XH-1 but decreased the abundance of strain LJD-29, indicating inhibitory effects of fungal preparation substrates on LJD-29 and promoting effects on XH-1 growth. The Bac treatment significantly increased the abundance of XH-1 (28.7%). Moreover, Fun, Bac, and Com treatments all increased the abundance of LJD-29, with similar effects observed in the Fun (86.2%) and Com (86.3%) treatments. Accordingly, after bioaugmentation, the relative abundance of strains added in the corresponding treatments increased markedly, indicating that bioaugmented strains LJD-20 and XH-1 persisted in these treatments.

To further explore the relations between added microorganisms and degradation efficiency, we analyzed the correlation between the PHE biodegradation and the abundance of added microorganisms. Figure 4 indicated significant positive correlations between the abundance of Aspergillus LJD-29 strain and PHE biodegradation (R² = 0.3691, P < 0.05), as well as between the abundance of Pseudomonas XH-1 strain and PHE biodegradation (R² = 0.5592, P < 0.05). In contrast, the abundance of Pseudomonas XH-1 exhibited a stronger correlation with PHE biodegradation.

Enzyme activities of fungi in soils

To investigate the mechanism of LJD-29 in situ degradation of PHE by indigenous fungi and bacteria, we assessed the enzyme activities of different soil treatments (Fig. 5). The results showed that lignin peroxidase activity increased significantly in the Fun and Com treatments, indicating that Aspergillus LJD-29 has a strong capacity to produce this enzyme during PHE degradation. Manganese peroxidase activity was consistently high across all treatments, suggesting a less pronounced role of Aspergillus LJD-29 in its production during PHE degradation.

Fig 5 — (a) Lignin peroxidase and (b) manganese peroxidase activities of the soil in each treatment after 8 days of incubation, respectively. (**c and d**) Pearson correlation coefficients between PHE degradation and enzyme activities in the Com treatment. NC, negative control using natural soil; NC-S, natural soil with fungal agent substrate; Bac, bioaugmentation with bacterial *Pseudomonas* XH-1; Fun, bioaugmentation with *Aspergillus* sp. LJD-29 agent; Com, bioaugmentation with both LJD-29 and XH-1; ST, sterilized soil.

Correlation analysis revealed significant positive associations between manganese peroxidase (R² = 0.2715, P < 0.05) and lignin peroxidase (R² = 0.4089, P < 0.05) activities with PHE biodegradation; lignin peroxidase activities showed a stronger correlation with PHE biodegradation. This finding highlighted the important roles of these enzymes in the PHE degradation process. Furthermore, our results of functional ASV identification indicated direct bacterial involvement in PHE degradation, while fungi mainly secreted the aforementioned enzymes to transform PHE. This supported the distinct biodegradation mechanisms of Aspergillus LJD-29 and Pseudomonas XH-1 in the tested soil.

Validation of fungal LJD-29 and bacterial XJH-1 in PHE degradation and assimilation

To elucidate the PHE co-degradation mechanism by LJD-29 and XH-1 and their roles in fungal–bacterial bioaugmentation, we conducted a validation experiment using pure cultures in mineral salts medium (MSM). Based on the results of SIP and PHE degradation analysis, strain LJD-29 removed 68.3% of PHE after 8 days of cultivation but did not directly assimilate PHE as a carbon source (Fig. S3 and S4). In contrast, bacterial XH-1 degraded 84.9% of PHE and demonstrated clear assimilation of PHE, as indicated by the significant enrichment of ¹³C-labeled DNA at higher BDs in the ¹³C-XH microcosm (Fig. 6). Notably, the addition of both LJD-29 and XH-1 (LJD-XH treatment) resulted in improved PHE removal efficiency (94.8%) compared to individual treatments, although the enhancement effect was not significant compared to theXH treatment (MSM with addition of strain XH-1). However, qPCR analysis of 16S rRNA genes demonstrated enhanced assimilation of bacterial XH-1 toward PHE in the LJD-XH treatment. The REF value of bacterial XH-1 in the LJD-XH treatment (12.1) was significantly higher than that in treatment with XH-1 alone (7.16; Fig. 6). Conversely, qPCR analysis of ITS genes indicated no fungal involvement in PHE assimilation during LJD-XH treatment. These findings suggested that while the simultaneous addition of fungi and bacteria may not substantially improve pollutant removal compared to individual treatments, it effectively enhanced the efficiency of the assimilation of organic pollutants by microorganisms.

Fig 6 — Relative abundances of 16S rRNA gene in CsCl gradient fractions of SIP experiment under pure culture conditions. (a) and (b) represent treatment with strain XH-1 and treatment with both strains LJD-29 and XH-1. The y-axes represent the gene relative abundance at each gradient fraction. (c) REF values of strain XH-1 in the treatments with strain XH-1 and treatment with both strains LJD-29 and XH-1. (d) REF values of strain LJD-29 in the treatments with strain LJD-29 and treatment with both strains LJD-29 and XH-1.

Gas chromatography-mass spectrometry (GC-MS) analysis of PHE biodegradation by Aspergillus LJD-29 and Pseudomonas XH-1 revealed unique metabolic pathways (Tables S3 to S5; Fig. S5). In pure fungal cultures, 2,3-dihydro-4(1H)-phenanthrenone emerged as a potential primary PHE product (Table S3). Conversely, bacterial strain XH-1 produced 1-hydroxy-2-naphthalenecarboxylic acid as an identified product (Table S4). Intriguingly, cocultures of LJD-29 and XH-1 exhibited three potential compounds, 1-hydroxy-2-naphthalenecarboxylic acid, 3,5-dimethoxy-phenol, and 5-methyl-2-(1-methylethyl)−1-hexanol (Table S5). These findings highlighted the distinct metabolic pathways used by the two microorganisms in PHE degradation. Fungi predominantly facilitate ring-opening of PHE pollutants, while bacteria directly convert pollutants into smaller molecules, actively engaging in the degradation of fungal metabolites.

DISCUSSION

Microorganisms, particularly fungi and bacteria, play vital roles in the degradation of organic pollutants in soil environments (23, 27). Indigenous bioaugmentation with fungi or bacteria has been shown to effectively enhance the removal of organic pollutants, primarily through changes in functional microorganisms or gene enrichment (13, 17, 21). However, the mechanisms underlying the combined degradation of organic pollutants by fungi and bacteria remain poorly understood. In this study, using PAHs, with a focus on PHE, as model compounds (28, 29), we used SIP and amplicon-sequencing techniques to elucidate the mechanism of fungal–bacterial co-degradation in petroleum-contaminated soil. Our results showed that bioaugmentation with both LJD-29 and XH-1 enhanced PHE degradation in PAH-contaminated soil, suggesting the potential of this mixed microbial community as a bioaugmentation remediation agent for PAH-contaminated soil. Each microbe played unique roles in the complex PAH degradation pathway by fungi and bacteria (30). In general, fungi use enzymes such as lignin and manganese peroxidases to convert PAHs into oxygenated intermediates, which are more hydrophilic and of higher reactivity, facilitating the subsequent degradation of PAHs by bacteria (17, 21, 31). In the present study, using SIP and metabolite analysis, we showed that fungi generated enzymes that initiate the breakdown of PHE without assimilating it. Conversely, bacteria not only enabled the comprehensive degradation of pollutants but also effectively degraded byproducts produced through fungal activity, resulting in significant enhancement of PHE assimilation efficiency.

SIP is a powerful tool for tracing functional microbial activity and has been used to investigate the degradation of soil organic pollutants and the bioaugmentation of indigenous bacteria (14, 32, 33). Here, we used SIP to explore the mechanism underlying fungal–bacterial co-degradation of PAHs. Interestingly, the results of SIP did not show the involvement of added fungi in PHE assimilation. However, this did not negate the potential role of fungi in in situ PHE biodegradation. Analysis of the correlations between enzyme activity and degradation efficiency suggested that the added fungus can facilitate PHE transformation through the secretion of manganese peroxidase and lignin peroxidase without directly assimilating PHE. Strikingly, no fungi were shown to be involved in pollutant assimilation in either bioaugmentation (Bac, Fun, and Com) or non-bioaugmentation (NC and NC-S) treatments. Fungi primarily contribute to the ring opening of PAHs, especially high molecular weight PAHs, and cannot fully remove PAH contaminants (1, 34). The mechanism of degradation involves the secretion of extracellular enzymes that break down the aromatic rings of PAHs into smaller molecules, facilitating their further uptake into microbial cells (35, 36). These enzymes, including laccase, manganese peroxidase, and lignin peroxidase, are components of the ligninolytic enzyme system and exhibit broad substrate specificity, particularly toward high molecular weight PAHs. Our findings emphasized the role of fungi in fungal–bacterial bioaugmentation, enhancing PAH degradation by transforming PAHs through extracellular enzymes and facilitating their assimilation by functional bacteria.

Bacteria, particularly Pseudomonas XH-1, play vital roles in PHE assimilation and degradation. Our SIP results confirmed the active participation of XH-1 in PHE assimilation across all bioaugmentation treatments. In addition, strain XH-1 exhibited a significant positive correlation with PHE degradation effectiveness, indicating its crucial role in PHE degradation. Moreover, XH-1 was shown to be an important indigenous PHE-degrading microorganism in the natural soil (NC), indicating its potential as a bioremediation agent for PAH-contaminated soil. Although efficient degrading bacteria are widely used for fungal–bacterial co-degradation of PAHs (30, 37), the dynamics of functional microbial communities and the roles of added microorganisms in this process have not been examined. This study provided compelling evidence of the role of the added bacterium in the co-degradation of PAHs by the fungus and bacterium. Pseudomonas sp., known for its ability to grow on various aromatic compounds including PAHs, possesses multiple dioxygenases and related enzymes involved in the degradation of aromatic compounds (38, 39). Using SIP, we successfully traced the in situ PHE degradation capability of this microorganism and discovered its potential for PAH bioremediation.

In addition to the enhancement of PHE degradation and assimilation, the added strains also modified the microbial community structure of other functional bacteria, resulting in improved PHE degradation efficiency. Our study demonstrated the impact of fungal–bacterial bioaugmentation on the microbial community structure of PAH-degrading bacteria in petroleum-contaminated soil and showed that unique microbial communities within a shared environment can lead to varying levels of PHE degradation activity. Our research illustrated this influence, and our observations have important implications for understanding the fungal–bacterial co-degradation mechanisms of PAHs. The natural soil harbored PAH-degrading bacterial functional microorganisms, such as Mycobacterium, Mycolicibacterium, and Ensifer, which are known for their ability to degrade PAHs (40 –42). Mycobacterium and Mycolicibacterium can use pyrene, PHE, and fluorene as carbon and energy sources and encode the alpha subunit of a PAH ring-hydroxylating dioxygenase (41, 43). However, with the addition of bacterial XH-1, other active functional bacteria from three additional genera (Nocardioides, Micromonospora, and Ramlibacter) were identified, along with the participation of XH-1 in PHE degradation. Nocardioides carries the genes for PHE dioxygenase, phdABCD, enabling the oxidation of PHE (40, 44). Micromonospora and Ramlibacter have also been reported to be associated with PAH degradation (45 –47). Our study further revealed the ability of these microbes to degrade PHE in situ. In the presence of the fungus LJD-29, Lysobacter emerged as a significant microorganism for in situ PHE degradation, and this microbe was reported to be able to efficiently use naphthalene and PHE as its sole carbon and energy sources (48). When fungal and bacterial microorganisms are simultaneously added for bioaugmentation, other functional bacteria, including Streptomyces, contribute to PHE degradation. Streptomyces, typically involved in the degradation of petroleum, possesses the genetic and physiological potential to degrade PHE (49).

It is worth noting that the combination of fungi and bacteria did not significantly enhance the degradation rate of PHE compared to treatment with bacteria alone; however, it did enhance the microbial assimilation of PHE. This suggested a synergistic role between fungi and bacteria in PAH degradation. In a single fungal and bacterial experimental system, fungi could use extracellular enzymes to transform PAHs, which cannot be fully removed, whereas bacteria could metabolize PAHs directly (34). However, the combined effects of fungi and bacteria in the degradation of PAH in soil environments remain unclear. The observed phenomena may be attributable to the unclear mechanism of interaction between fungi and bacteria, their growth conditions (50, 51), and elevated levels of inhibitory metabolites. Numerous studies have reported that biological metabolites of PHE have the potential to impede microbial degradation of PHE (52, 53). Nevertheless, our study demonstrated significant enhancement of the bacterial assimilation of PHE, suggesting that bacteria play crucial roles in converting PAHs into biomass or metabolites. We showed that the synergistic effect between fungi and bacteria did not directly increase the PAH degradation efficiency but substantially improved microbial assimilation. This is important in understanding the interaction and mechanisms involved in the elimination of PAHs by fungi and bacteria.

In this study, we used bioaugmentation of indigenous fungus LJD-29 and bacterium XH-1 to enhance the biodegradation of PAH in petroleum-contaminated soil. The synergistic action of fungi and bacteria not only increased the efficiency of PHE biodegradation but also greatly enhanced PHE assimilation efficiency. The fungus and bacterium played distinct roles in the degradation process, demonstrating their collaborative involvement. The fungus used extracellular enzymes to partially transform PHE compounds, but complete degradation was not achieved, whereas the bacterium actively participated in the transformation and assimilation of PHE. The results of SIP experiments under pure culture conditions further confirmed these findings. Moreover, the added fungus and bacterium both played crucial roles in PHE degradation in natural soil, as demonstrated by SIP and correlation analyses. These observations underscored the significance of enhancing indigenous functional microorganisms to expedite the degradation and assimilation of PAHs and other organic pollutants in contaminated soil. In addition, the bioaugmentation by these microorganisms altered the community structure of soil PHE-degrading bacteria. Overall, this study revealed a novel mechanism of fungal–bacterial co-bioremediation, provided insights into the diversity of PHE-degrading communities, and proposed indigenous fungal–bacterial biodegradation as a promising in situ treatment strategy for PAH-contaminated soil.

MATERIALS AND METHODS

Sample collection

Soil samples were collected from the surface layer (0–10 cm) of a petroleum refinery in Shandong Province, China (37°46′N, 118°50′E). The samples were sieved through a 5 mm mesh and stored in the dark at 4°C. Some samples were stored at −80°C for DNA extraction, while others were reserved for bacterial–fungal co-biodegradation experiments. The collected soil had the following physicochemical properties: pH 7.6; organic matter content, 10.4 g·kg⁻¹; total nitrogen content, 0.72 g·kg⁻¹; total phosphorus content, 0.59 g·kg⁻¹; and total petroleum hydrocarbon content, 1.19 g·kg⁻¹.

Cultivation and preparation of fungal and bacterial strains

The fungal strain Aspergillus sp. LJD-29 and the bacterial strain Pseudomonas XH-1 were enriched and isolated from petroleum-contaminated soil using MSM supplemented with PHE (50 mg·L⁻¹) as the sole carbon source. Isolation was performed using the spread plate method on MSM agar medium. Detailed information regarding this process is available in the Supporting Information. After 8 days of incubation in MSM, strains LJD-29 and XH-1 degraded over 65% of PHE from an initial concentration of 50 mg·L⁻¹ (Fig. S1).

The fungus and its agent were prepared as described in our previous report (17). Briefly, the fungal agent substrate, consisting of a 9:1 (wt/wt) mixture of pine sawdust/wheat bran and corn starch as a binder, was sterilized by autoclaving at 121°C for 25 min. The activated fungal culture (Aspergillus sp. LJD-29) was transferred to 250 mL conical flasks containing 100 mL of MSM and incubated in a shaking incubator at 28°C and 150 rpm for 7 days. The harvested fungal mycelia were washed, weighed, and adjusted to a concentration of approximately 10 g·L⁻¹ with sterile water. The mycelia were then disrupted using a tissue homogenizer on the low-speed setting to obtain a suitable fungal suspension. This fungal suspension, mixed with 3% sodium alginate, was added to the prepared substrate. The inoculated substrates were precultivated at 28°C for 7 days until the surface was covered with fungal mycelia.

Strain XH-1 was cultivated in the dark at 30°C in MSM supplemented with 50 mg·L⁻¹ PHE for 4 days with shaking at 180 rpm. Enriched cultures were transferred to sterile centrifuge tubes, and pelleted cells were obtained after centrifugation at 4,000 × g for 10 min. The cells were then resuspended in phosphate-buffered saline (PBS). The population of strain XH-1 was determined to be approximately 5 × 10⁹ CFU/mL using the dilution plate counting method (54, 55).

SIP microcosm for fungal–bacterial biodegradation

Microcosms were established in seeding trays (5.4 × 5.8 × 5.0 cm) containing 5 g of sampled soil (dry weight) with or without microorganism inoculation. For bacterial strain treatments, 100 µL of XH-1 cell/PBS suspension was added, resulting in an initial population of approximately 1 × 10⁸ CFU/mL. Fungal strain treatments consisted of 15% [(wt/wt) dry basis] pregrown inoculum on a substrate of strain LJD-29. Combined treatments included 15% [(wt/wt) dry basis] of strain LJD-29 fungal agent and 100 µL of bacterial cell/PBS suspension. The unlabeled PHE (99%) or ¹³C-labeled PHE (¹³C₁₄-PHE, 99%; Cambridge Isotope Laboratories, Inc., Tewksbury, MA, USA) was added to the above trays at a final concentration of 5 mg·L⁻¹. In this study, the biotic treatments were as follows: ¹²C_NC (¹²C-PHE without any strains), ¹²C_NC-S (¹²C-PHE with fungal agent substrate), ¹²C_Fun (¹²C-PHE with fungal strain LJD-29), ¹²C_Bac (¹²C-PHE with bacterial strain XH-1), ¹²C_Com (¹²C-PHE with both strains LJD-29 and XH-1), ¹³C_NC (¹³C-PHE without any strains), ¹³C_NC-S (¹³C-PHE with fungal agent substrate), ¹³C_Fun (¹³C-PHE with fungal strain LJD-29), ¹³C_Bac (¹³C-PHE with bacterial strain XH-1), and ¹³C_Com (¹³C-PHE with both strains LJD-29 and XH-1). Microcosms with unlabeled PHE in ST were also established. Soil moisture was adjusted to 60% water-holding capacity, and the microcosms were incubated as described previously (17). Destructive sampling was conducted on days 2, 4, 6, and 8, with freeze-dried and sieved samples stored at −20°C for subsequent chemical analysis. DNA extraction was performed on day 8 for all treatments considering poor PHE degradation efficiencies in some treatments, such as NC and NC-S.

DNA extraction, ultracentrifugation, qPCR, and amplicon sequencing

DNA extraction was conducted using our previously described methods on different microcosms from each treatment (22, 56). CsCl-gradient ultracentrifugation was performed to separate DNA from the ¹²C- and ¹³C-PHE microcosms. Subsequently, bacterial 16S rRNA genes in all DNA fractions from the microcosms were amplified. The DNA fractions with BD values of 1.6914–1.6929 g·m⁻¹ and 1.7019–1.7089 g·m⁻¹ were identified as the light and heavy DNA (¹³C-DNA) fractions, respectively, for the ¹³C_PHE microcosms, based on the relation between BD and 16S rRNA gene abundance (Fig. S2). Further details are presented in the Supporting Information.

For amplicon sequencing, the hypervariable V4 region and ITS region were amplified from the total DNA extracted from all treatments and the separated DNA from the SIP treatments. Primer sets 515F/806R and ITS3F/ITS4R (Table S1) were used for amplification, with a 12 bp-specific barcode inserted into the pre-primer for sample labeling and identification in downstream bioinformatics analysis (17, 57, 58). The PCR mixtures (25 µL) consisted of 12.5 µL of TaKaRa rTaq premixed buffer, 0.5 µL of primers at 100 nM, 1 µL of template DNA, and 10.5 µL of ddH₂O. PCR consisted of 35 cycles of denaturation at 95°C for 60 s followed by annealing at 55°C for 45 s and extension at 72°C for 60 s. Amplified products from the same sample were merged, purified, and quantified using a Qubit 2.0 fluorometer (59). Equal amounts of DNA from different sources were combined, further purified by agarose gel electrophoresis, and recovered using an Agarose Gel DNA Fragment Recovery Kit (60). DNA concentration was measured, and the samples were submitted to the Beijing Genomics Institute (BGI) for sequencing.

Computational analyses for PHE degrader identification

The raw sequence data from Illumina MiSeq were assembled and analyzed using QIIME 2 (v. 2020.2). Primer sequences were removed from the reads using Cutadapt (v. 2018.11.0), and the ITS2 reads were preprocessed using ITSxpress (v. 1.7.2) (61). Quality control, chimera filtering, noise reduction, and merging were performed using the DADA2 pipeline to generate bacterial and fungal ASVs with 100% similarity (62). Bacterial and fungal ASVs were categorized using SILVA 132 and UNITE reference databases (63).

To identify functional microorganisms involved in the in situ PHE degradation, we used a method based on REF as described previously (22, 23). The REF was calculated using the formula REF = (13C_heavy/13C_light)/(12C_heavy/12C_light), where 13C_Heavy and 13C_Light represent the relative abundances of ASVs in the heavy and light DNA fractions of the ¹³C-PHE treatment, and 12C_Heavy and 12C_Light represent the relative abundances of ASVs in the heavy and light DNA fractions of the ¹²C-PHE treatment, respectively. When the REF value exceeded 1.5, microorganisms represented by ASVs were considered active functional bacteria or fungi. Active degraders were identified by calculating the REF value for ASVs with a relative abundance (>0.1%) among the top 300 ASVs in different treatments. Phylogenetic analysis of the identified sequences was performed as described previously (9, 64).

Chemical and enzymatic analyses

PHE in soil was extracted from each treatment containing unlabeled PHE on days 2, 4, 6, and 8 of incubation. Analysis was performed by GC-MS (Agilent 7890; Agilent Technologies, Santa Clara, CA, USA) according to the methods described previously (10). Soil samples were prepared, spiked with recovery standards (deuterated PHE), and extracted using dichloromethane. The organic extract was then purified using a silica gel/alumina column and concentrated to approximately 0.5 mL. Prior to instrumental analysis, 1,000 ng of hexamethylbenzene was added to the organic solvents as an internal standard.

As strain LJD-29 can produce lignin peroxidase and manganese peroxidase (63), these two peroxidases were measured using an ultraviolet spectrophotometer following the instructions provided with the test kits (Beijing Solarbio Science and Technology Co., Ltd., Beijing, China). Lignin peroxidase activity was determined by measuring the absorbance at 310 nm, which corresponds to the oxidation of veratrol. Manganese peroxidase activity was determined by measuring the absorbance at 465 nm, which corresponds to the formation of tetra-o-methoxycophenol.

SIP experiment under pure culture conditions and PHE metabolite analysis

To further investigate the co-biodegradation mechanism of strains LJD-29 and XH-1 and their roles in fungal–bacterial bioaugmentation, we conducted SIP experiments and measured the degradation efficiency of the selected strains using MSM culture. A total of six treatments, i.e., ¹²C-LJD (MSM with ¹²C-PHE and strain LJD-29), ¹²C-XH (MSM with ¹²C-PHE and strain XH-1), ¹²C-LJD + XH (MSM with ¹²C-PHE and both strains), ¹³C-LJD (MSM with ¹³C-PHE and strain LJD-29), ¹³C-XH (MSM with ¹³C-PHE and strain XH-1), and ¹³C-LJD + XH (MSM with ¹³C-PHE and both strains), were used. After completing the cultivation, DNA extraction, ultracentrifugation, and qPCR (as described in Materials and Methods) were performed to calculate the REF of strains and assess their PHE assimilation effectiveness.

Moreover, to examine the mechanism of fungal–bacterial PHE metabolism further, we conducted a comprehensive analysis of metabolites across different treatments. Following incubation, 100 mL of ethyl acetate was added with shaking at 160 rpm for 2 h for metabolite extraction. Subsequently, the combined extraction liquids were concentrated by rotary evaporation under vacuum, and the residual solution was dried under a gentle stream of nitrogen. The resulting residue was dissolved in 0.5 mL of hexane. PHE metabolites were identified by GC-MS using an Agilent 7890 instrument programmed to increase in temperature from 80°C to 260°C at a rate of 5°C min⁻¹, with a final hold time of 27 min.

Statistical analysis

Data were analyzed by one-way analysis of variance followed by Tukey’s post hoc test for multiple comparisons. Pearson’s correlation analysis was performed to examine the relations between degradation rates and influencing factors, with a significance criterion of α = 0.05. The phylogenetic information of ASV was obtained using the Basic Local Alignment Search Tool (BLAST) and MEGA version 11. Phylogenetic trees were constructed using the neighbor-joining method with Jukes Cantor correction, and support was calculated with 1,000 repeated resamplings. All statistical analyses were performed using SPSS 19.0 (SPSS Inc., Chicago, IL, USA).

ACKNOWLEDGMENTS

This work was supported by the National Natural Science Foundation of China (Nos. 32061133003 and 42277210), Natural Science Foundation of Guangdong Province, China (2023B1515020038), Guangdong Foundation for Program of Science and Technology Research (2023B1212060049), and the Youth Innovation Promotion Association CAS (2023368).

We acknowledge the cooperation between China and the EU through the EiCLaR project (European Union’s Horizon 2020, N° 965945).

Contributor Information

Jibing Li, Email: lijibing@gig.ac.cn.

Irina S. Druzhinina, Royal Botanic Gardens, Surrey, United Kingdom

DATA AVAILABILITY

Sequence data have been deposited in the National Center for Biotechnology Information database under the accession number PRJNA1089363.

SUPPLEMENTAL MATERIAL

The following material is available online at https://doi.org/10.1128/aem.00662-24.

Supplemental material. aem.00662-24-s0001.doc.

Tables S1 to S5; Figures S1 to S5.

aem.00662-24-s0001.doc^{(1MB, doc)}

DOI: 10.1128/aem.00662-24.SuF1

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Associated Data

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Supplementary Materials

Supplemental material. aem.00662-24-s0001.doc.

Tables S1 to S5; Figures S1 to S5.

aem.00662-24-s0001.doc^{(1MB, doc)}

DOI: 10.1128/aem.00662-24.SuF1

Data Availability Statement

Sequence data have been deposited in the National Center for Biotechnology Information database under the accession number PRJNA1089363.

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PERMALINK

Determination of soil phenanthrene degradation through a fungal–bacterial consortium

Chunling Luo

Guoqing Guan

Yeliang Dai

Xixi Cai

Qihui Huang

Jibing Li

Gan Zhang

Roles

ABSTRACT

IMPORTANCE

INTRODUCTION

RESULTS

PHE degradation

Fig 1.

PHE-degrading bacteria and fungi revealed by DNA-SIP

Fig 2.

Fig 3.

The abundance of Aspergillus LJD-29 and Pseudomonas XH-1 and their effects on PHE degradation

Fig 4.

Enzyme activities of fungi in soils

Fig 5.

Validation of fungal LJD-29 and bacterial XJH-1 in PHE degradation and assimilation

Fig 6.

DISCUSSION

MATERIALS AND METHODS

Sample collection

Cultivation and preparation of fungal and bacterial strains

SIP microcosm for fungal–bacterial biodegradation

DNA extraction, ultracentrifugation, qPCR, and amplicon sequencing

Computational analyses for PHE degrader identification

Chemical and enzymatic analyses

SIP experiment under pure culture conditions and PHE metabolite analysis

Statistical analysis

ACKNOWLEDGMENTS

Contributor Information

DATA AVAILABILITY

SUPPLEMENTAL MATERIAL

REFERENCES

Associated Data

Supplementary Materials

Data Availability Statement

ACTIONS

PERMALINK

RESOURCES

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