Effects of copper on expression of methane monooxygenases, trichloroethylene degradation, and community structure in methanotrophic consortia

Abstract

Copper plays a key role in regulating the expression of enzymes that promote biodegradation of contaminants in methanotrophic consortia (MC). Here, we utilized MC isolated from landfill cover to investigate cometabolic degradation of trichloroethylene (TCE) at nine different copper (Cu²⁺) concentrations. The results demonstrated that an increase in Cu²⁺ concentration from 0 to 15 μM altered the specific first‐order rate constant k _1,TCE, the expression levels of methane monooxygenase (pmoA and mmoX) genes, and the specific activity of soluble methane monooxygenase (sMMO). High efficiency TCE degradation (95%) and the expression levels of methane monooxygenase (MMO) were detected at a Cu²⁺ concentration of 0.03 μM. Notably, sMMO‐specific activity ranged from 74.41 nmol/(mg_cell h) in 15 μM Cu²⁺ to 654.99 nmol/(mg_cell h) in 0.03 μM Cu²⁺, which contrasts with cultures of pure methanotrophs in which sMMO activity is depressed at high Cu²⁺ concentrations, indicating a special regulatory role for Cu²⁺ in MC. The results of MiSeq pyrosequencing indicated that higher Cu²⁺ concentrations stimulated the growth of methanotrophic microorganisms in MC. These findings have important implications for the elucidation of copper‐mediated regulatory mechanisms in MC.

Abbreviations

MC

methanotrophic consortia

MMOs

methane monooxygenases

MNMS

modified nitrate mineral salts

NMS

nitrate mineral salts

OTUs

Operational Taxonomic Units

pMMO

particulate methane monooxygenase

sMMO

soluble methane monooxygenase

TCE

trichloroethylene

VOCs

volatile organic compounds

1. Introduction

Trichloroethylene (TCE) is a frequently found soil and groundwater contaminant that results from its widespread use in various industrial processes and improper disposal methods 1, 2. Since TCE is toxic to humans, a possible carcinogen, and recalcitrant 3, 4, this is a serious environmental concern. TCE is also emitted, along with methane (CH₄), by anaerobic degradation of organic compounds present in landfills 5, 6. There have been several reports concerning the co‐oxidization of TCE using CH₄ as a substrate by landfill cover soil, and subsequent results suggests that the methanotrophic bacteria in landfill cover soil plays an important role in degradation of CH₄ and TCE 7, 8.

Over the years, studies have focused on the cometabolic degradation of TCE by methanotrophs. The key enzyme for CH₄ and TCE degradation in methanotrophic bacterial is CH₄ monooxygenase (MMO). MMO is present in two forms: membrane‐associated or particulate methane monooxygenase (pMMO), which is found in most known methanotrophs and is localized to the cytoplasmic membrane, and soluble methane monooxygenase (sMMO). sMMO is expressed in only some methanotrophs (e.g. Methylosinus, Methylococcus) and is the only MMO expressed in Methylocella 9, 10 and Methyloferrula 11. pMMO and sMMO are encoded by the genes pmoA and mmoX, respectively 12.

It is known that copper (Cu²⁺) plays a key role in regulating the expression of MMOs as well as the activity of these enzymes 13, 14, 15. Prior results suggest that a high copper‐to‐biomass ratio stimulates pMMO and represses sMMO expression in pure cultures 13, 16, 17. Furthermore, Cu²⁺ has been shown to alter pMMO expression up to 55‐fold and to change its substrate affinity and specificity 18, 19, 20. It has also been reported that sMMO activity is undetectable at a copper‐to‐biomass ratio of 45.64 mmol Cu·g⁻¹ _protein in the culture medium of Methylosinus trichosporium OB3b 21. Thus, Cu²⁺ appears to function as a switch regulating MMO expression and highlights the likely role of Cu²⁺ in regulating biodegradation of CH₄ and TCE 15.

Prior studies have explored the effects of Cu²⁺ on CH₄ oxidation and TCE biodegradation. Smith et al. 22 studied the influence of Cu²⁺ on the kinetic parameters of CH₄ and TCE in the methanotroph Methylobacter sp. strain BB5.1 grown on medium. The results suggested that cells grown on low‐copper medium did not oxidize TCE and had a variable rate of CH₄ oxidation. Similarly, Sonny Lontoh et al. 19 explored the effects of Cu²⁺ on CH₄ and TCE degradation and enzyme expression using the model methanotroph Methylosinus trichosporium OB3B. The results indicated that the kinetic parameters of CH₄ oxidation decreased along with Cu²⁺ concentrations (2.5–20 μM) and the affinity for TCE increased with increasing Cu²⁺ by providing formate only when expressing pMMO. Subsequently, many researchers have examined the activity of pMMO and sMMO in chlorinated ethene degradation by changing Cu²⁺ concentrations to regulate MMO expression 23, 24, 25, 26. As a result, it was suggested that expression and activity of sMMO was essential for effective degradation of chlorinated hydrocarbons and that sMMO had a broader specificity than pMMO in regard to degradation of chlorinated hydrocarbons.

Environmental microbiology is faced with an enormous diversity of organisms that remain individually uncultured, and microbial communities have demonstrated a significant advantage over pure strains in regard to their application toward biodegradation 27, 28. Some bacteria have evolved to utilize a variety of pollutants; however, the complete degradation of pollutants typically depends on the collective interaction of a diverse assortment of microorganisms, with any individual microorganism not able to fully complete the degradation process 29. Since most work regarding biodegradation relies on studies using pure cultures, the effects of Cu²⁺ on the degradation of pollutants, genes encoding key enzymes, and changes to the microbiology community in MC remains largely unexplored. Thus, in this study, the cometabolic degradation of TCE with CH₄ substrate at nine different Cu²⁺ concentrations was assessed in MC. Furthermore, the expression of pmoA and mmoX was quantified by qRT‐PCR under various conditions to assess the expression of this key enzyme. Finally, the structure of the microbe community under different Cu²⁺ concentrations was investigated via MiSeq sequencing in order to explore the diversity of this functional population.

2. Materials and methods

2.1. Enrichment of MC

To obtain MC, landfill cover soil was sampled at Chang Shengqiao municipal solid‐waste landfill, Chongqing, China. Detailed information regarding the landfill was previously described by Xing Zhilin et al. 30. Landfill cover soil was sampled from 30 cm below the surface by extracting cores, passed through a 2 mm‐sieve to remove the large particles, and then stored on ice. A total of 100 g (wet weight) of soil sample was added to a 500 mL serum bottle and then sealed with butyl rubber and aluminum cap. Subsequently, 100 mL of CH₄ gas (99.999%; Chongqing Ruixin Gases Co., China) was injected into the headspace of the bottle via a gas‐tight syringe to introduce an initial concentration of 20% CH₄. The bottle was then incubated at 30°C. The CH₄ concentrations in the headspace were periodically measured before and after incubation. When CH₄ concentrations dropped below the detection limit, the serum bottle was opened for 1 h to allow the gases inside the bottle to be replaced with air. The serum bottle was then resealed with a butyl rubber stopper and CH₄ was reinjected at the same concentration as described above. The bottle was then re‐incubated at 30°C. This process lasted 2 weeks. The microbial activity of the enriched landfill cover soil was investigated via batch experiments. The experimental method was performed as previously described 30. The biodegradation of CH₄ and TCE are shown in Supporting Information Fig. 1. CH₄ (10% V/V) was completely degraded after 7 days, and 75% of TCE (20 μM) was removed after 3 days.

A total of 2 g (wet weight) of the above enriched soil was then added to a 100 mL serum bottle with 20 mL modified nitrate mineral salts (MNMS, without Cu²⁺) medium 31 and then mixed by vortexing for 3 min and placed on a shaker platform at 170 rpm for 2 h. Subsequently, 1 mL of supernatant obtained from the serum bottle was added to a fresh 100 mL serum bottle with 20 mL nitrate mineral salts (NMS) medium. The serum bottle was then sealed with an aluminum crimp cap and 20 mL CH₄ gas was injected to replace the equivalent volume of air. The caps contained a butyl rubber septum to enable gas sampling from the headspace using a gas‐tight syringe. The bottle was then incubated at 30°C at 170 rpm until the cell subculture was at the exponential stage.

All experiments were performed in triplicate. A 1 mL CH₄ sample was withdrawn with a syringe with a Teflon plunger (Agilent, USA) from each serum bottle and directly analyzed using a thermal conductivity detector on a SC‐6000A gas chromatograph (Chuanyi Spectrometry Instruments, Chongqing, China). The temperature profiles and parameters used were as follows: injector, 120°C; column oven, 90°C; detector, 120°C; carrier gas, and N₂ at 25 mL/min.

2.2. Cometabolic degradation of TCE at different concentration of Cu²⁺

The MC was initially grown to the mid‐exponential phase on MNMS at 30°C and 170 rpm in a CH₄‐to‐air ratio of 1:2. Biomass was harvested by centrifuging at 7818 × g for 5 min and then washed twice with fresh MNMS medium to remove residual CH₄. The cell pellet was resuspended in fresh MNMS medium to an OD₆₀₀ of 0.1. TCE degradation was then performed in 100 mL serum bottles with 20 mL resuspended cell culture at different Cu²⁺ concentrations achieved by adding CuSO₄. The range of Cu²⁺ concentrations used in this study was 0–15 uM. The serum bottles were sealed with Teflon‐lined butyl rubber stoppers, and 15 mL of the headspace was then replaced with CH₄ using a gas‐tight syringe to achieve a final concentration of about 15% v/v. TCE was then added to achieve an initial aqueous concentration of 6 μM. All experiments were conducted with duplicate biological replicates. All serum bottles were incubated at 30°C and 170 rpm.

The headspace biogas and TCE were assessed every day via gas chromatography with an electron capture detector. Injector, oven, and detector temperatures were set to 200, 80, and 250°C, respectively. The N₂ carrier gas flow rate was set to 40 mL/min.

2.3. RNA extraction and quantitative analysis of MC using qRT‐PCR

The transcriptional expression levels of pmoA and mmoX in the test samples under different Cu²⁺ concentrations were quantified using reverse transcription PCR and real‐time quantitative PCR. RNA was extracted from the stored cell pellets using an RNAprep mini kit (Tiangen Bio‐Chem Technology, Beijing, China) and quantified using a DeNovix spectrophotometer (DeNovix, USA) at 260 nm. cDNA was then synthesized from the purified total RNA extracts using a PrimeScript™ RT reagent kit with gDNA Eraser (Perfect Real Time) by RT‐PCR, following the manufacturer's instructions (Takara Bio Technology, Dalian, China). The gel electrophoresis images of the RNA and cDNA are shown in Supporting Information Fig. 2. RT‐qPCR was performed on a CFX48 real‐time PCR system (Bio‐Rad, USA) with SYBR® Premix Ex Taq™ (Tli RNaseH Plus; Takara Bio Technology, Dalian, China) using primers previously published 32 and those designed using Primers 5 (Supporting Information Table 1). The transcript levels of each functional gene were normalized against the 16S rRNA gene copy number as quantified by qPCR. The copy numbers of the 16S rRNA, pmoA, and mmoX transcripts were calculated from measured Ct values using a calibration curve based on seven plasmid preparations with known copy numbers. Additional details regarding RNA and qPCR assays are provided in the Supporting Information.

The qRT‐PCR program was as follows: initial denaturation for 30 s at 95°C followed by 39 cycles consisting of denaturation at 95°C for 5 s, primer annealing at 62.4°C (for 16SrRNA/pmoA gene) or 58°C (for mmoX gene) for 30 s, and melting at 65 to 95°C with increments of 0.5°C.

2.4. DNA extraction and sequencing

The bacterial community structure was assessed by MiSeq sequencing using the adapted primers 338F (ACTCCTACGGGAGGCAGCA) and 806R (GGACTACHVGGGTWTCTAAT) for the V3‐V4 regions of the 16S rRNA gene 33. PCR amplification was performed as previously published 34. Equal quantities of the three PCR products per sample were pooled and purified using a QIAquick PCR purification kit (Qiagen, USA). Finally, a mixture of the amplicons from the different samples was sent to the MiSeq Illumina platform at Shanghai Major Bio‐pharm Biotechnology Co. (Shanghai, China) for sequencing. The treatment of the raw sequences is detailed in the Supporting Information. One‐way analysis of variance (ANOVA) was used to assess the homogeneity of the chemical parameters of the cover soil with a significance level of 5% (p < 0.05).

2.5. Kinetics of cometabolic degradation

At low substrate concentrations, where the substrate concentration C << half saturation constant of the substrates K _m, the transformation follows first‐order kinetics. In this case, the microbial degradation of the substrate can be described as:

$$-\frac{\mathrm{d}C}{\mathrm{d}t}=kXC$$ (1)

Where C is the substrate (CH₄ or TCE) concentration in a phase, t is the reaction time, k is first‐order rate constant of the substrate, and X is the cell concentration. In this study, the CH₄ concentration in the gaseous phase and TCE concentration in the liquid phase were used in Eq. (1). As such, the first‐order kinetics equation of CH₄ and TCE can be described with (2) and (3), respectively:

$$-\frac{\mathrm{d}{C}_{\mathrm{g},\mathrm{C}{\mathrm{H}}_4}}{\mathrm{d}t}=k_{g,C{H}_4}X{C}_{\mathrm{g},\mathrm{C}{\mathrm{H}}_4}$$ (2)

$$-\frac{\mathrm{d}{C}_{\mathrm{l},\mathrm{T}CE}}{\mathrm{d}t}=k_{\mathrm{l},\mathrm{TCE}}X{C}_{\mathrm{l},\mathrm{TCE}}$$ (3)

After integrating over time, these can be rewritten as:

$$C_{\mathrm{g},\mathrm{C}{\mathrm{H}}_4,\mathrm{t}}=C_{\mathrm{g},\mathrm{C}{\mathrm{H}}_4,0}{\mathrm{e}}^{-k_{\mathrm{g},\mathrm{C}{\mathrm{H}}_4}Xt}$$ (4)

$$C_{\mathrm{l},\mathrm{TCE},\mathrm{t}}=C_{\mathrm{l},\mathrm{TCE},0}{\mathrm{e}}^{-k_{\mathrm{l},\mathrm{TCE}}Xt}$$ (5)

Based on the relationship between time and substrate concentration in the system, the $k_{\mathrm{g},\mathrm{C}{\mathrm{H}}_4}$ mL/(mg_cells min) and k _{l, TCE} mL/(mg_cells min) can be calculated with Eqs. (4) and (5), respectively.

3. Results and discussion

3.1. Effect of Cu²⁺ on CH₄ oxidation and TCE degradation

The results of TCE degradation and the fitted curve with first‐order kinetics are shown in Fig. 1. The TCE degradation efficiency under the different Cu²⁺ concentrations ranged from 66.59% at 0.75 μM Cu²⁺ to 95.75% at 0.03 μM Cu²⁺ after 6 days. Furthermore, TCE and CH₄ transformation were well fitted by first‐order kinetics (R ² ranged from 0.81 to 0.95). The kinetic coefficients are shown in Supporting Information Table 2, the changes in the first‐order kinetic constants for TCE and CH₄ relative to Cu²⁺ concentration are shown in Fig. 2. The specific first‐order rate constants TCE k _1,TCE and CH₄ ${\mathrm{k}}_{\mathrm{g},\mathrm{C}{\mathrm{H}}_4}$ ranged from 0.044 to 0.075 and 0.051 to 0.124 mL/(mg_cells min), respectively. The specific first‐order rate constant of CH₄ was higher than that of TCE, which demonstrated that this key enzyme had a higher affinity for CH₄ relative to TCE in MC due to competitive inhibition 4, 35.

Figure 1.

First‐order kinetics of TCE degradation at different Cu²⁺ concentrations. Cu²⁺ concentrations in A, B, C, D, E, F, G, H and I were 0, 0.03, 0.2, 0.75, 1, 3, 5, 10 and 15 μM, respectively.

Figure 2.

Changes in first‐order kinetic constants of TCE and CH₄ relative to Cu²⁺ concentration.

In prior studies, Hylckama et al. 36 used Methylosinus trichosporium OB3b to assess the transformation kinetics of chlorinated ethenes and obtained a k _1,TCE of <0.03 mL/(mg_cells min) when Cu²⁺ was omitted from the medium. Similarly, Lontoh et al. 19 evaluated the cometabolic degradation of TCE using the same methanotroph and found that the k _{1, TCE} was 0.035 mL/(mg_cells min) when expression of sMMO was repressed by adjusting the Cu²⁺ concentration to 2.5 μM. Subsequently, Aziz et al. 37 assessed the cometabolism of chlorinated solvents using a mutant methanotroph Methylosinus trichosporium OB3b PP358 and determined the k _{1, TCE} was 0.014 mL/(mg_cells min). In this current study, the value of k _1,TCE ranged from 0.044 to 0.075 mL/(mg_cells min), which was higher than that previously reported using pure cultures, suggesting that MC had an substantial advantage over pure cultures 27, 28. Unexpectedly, k _{1, TCE} and ${\mathrm{k}}_{\mathrm{g},\mathrm{C}{\mathrm{H}}_4}$ fluctuated with increasing Cu²⁺ concentrations (0‐15 μM) in the cultures. The maximum of k _{1, TCE} occurred at Cu²⁺ concentrations of 0.03 μM and 5 μM by 0.064 and 0.075 mL/(mg_cells min), respectively. Similarly, the maximum of ${\mathrm{k}}_{\mathrm{g},\mathrm{C}{\mathrm{H}}_4}$ occurred at Cu²⁺ concentrations of 1 μM and 5 μM by 0.121 and 0.124 mL/(mg_cells min), respectively. These observations indicated that the Cu²⁺ concentrations of 0.03 μM and 5 μM stimulated TCE degradation. These results were in agreement with the study of Adrian et al. 14, which indicated that a Cu²⁺ concentration of 2.5 μM demonstrated the highest TCE degradation within the range of 0–10 μM Cu²⁺ in a slurry incubation. It was reported that this stimulation was most likely caused by an increase of cell‐specific activity at low Cu²⁺ concentration. The stimulatory effect was diminished at >5 μM Cu²⁺, suggesting an adverse effect possibly resulting from copper toxicity 14.

3.2. Expression of pmoA and mmoX at different Cu²⁺ concentrations

Real‐time quantitative RT‐PCR was performed to confirm and quantify pmoA and mmoX expression in MC under different Cu²⁺ concentrations (Supporting Information Table 3). As shown in Fig. 3, the pmoA and mmoX gene copy number, which was normalized to overall rRNA transcript levels, ranged from 5.15 × 10⁻⁵ (4.63 × 10⁻⁶) at 3 μM Cu²⁺ to 4.22 × 10⁻³ (4.98 × 10⁻⁵) at 0.03 μM Cu²⁺ and 1.05 × 10⁻⁷ (3.39 × 10⁻⁸) at 3 μM Cu²⁺ to 9.30 × 10⁻⁶ (4.89 × 10⁻⁷) at 0.03 μM Cu²⁺, respectively. The expression level of pmoA was consistent with the study of Yoon et al. 24, which ranged from 1.69 × 10⁻³ (4.89 × 10⁻⁴) to 2.06 × 10⁻³ (6.23 × 10⁻⁴) in the Methylocystis strain SB2 at a Cu²⁺ concentration of 10 μM. Based on the expression of pmoA and mmoX, the quantity of pmoA was 1 to 4 orders of magnitude greater than that of mmoX, indicating that pMMO likely plays a more prominent role than sMMO under these conditions.

Figure 3.

Changes in abundance of target genes (pmoA/16S and mmoX/16S) relative to Cu²⁺ concentration during TCE biodegradation.

Prior research has shown that the expression of MMOs is regulated by Cu²⁺ 13, 16, 17. To be more accurate, the expression of sMMO is efficient at less than 0.8 μM Cu²⁺ and is restricted at an excess of 4 μM of Cu²⁺ in methanotrophs 38. In this current study, sMMO expression was not repressed at high Cu²⁺ concentrations (>4 μM), suggesting endosymbiotic interactions between methanotrophs and heterotrophs in MC. Numerous studies have shown that, as a heavy metal and coenzyme factor, Cu²⁺ can affect microbial activity, including microbial activity in activated sludge as well as methanotrophs 39, 40, 41. It can be speculated that other heterotrophic bacteria also feed on Cu²⁺ in MC, so the available concentration of Cu²⁺ for methanotrophs could be much less than the presumed concentration of Cu²⁺.

As shown in Fig. 4, the first‐order kinetic constants changed with the relative expression abundance of pmoA (Fig. 4A) and mmoX (Fig. 4B). As the expression of pmoA (0‐1.49 × 10⁻³) and mmoX (0‐5.02 × 10⁻⁶) increased, k _{1, TCE} increased substantially, while $k_{\mathrm{g},\mathrm{C}{\mathrm{H}}_4}$ decreased, which indicated a competitive inhibition between CH₄ and TCE in these cultures. During cometabolic degradation, competitive inhibition between the growth substrate and contaminants is unavoidable, as shown in prior studies 37, 42. Furthermore, the variation in the expression of the MMOs (pMMO and sMMO) and k _{1, TCE} at Cu²⁺ concentrations of 0–3 μM were the same, demonstrating that MMOs were the likely the major enzyme responsible for TCE degradation. This also supports the hypothesis that methanotrophs were the dominant microorganisms involved in this process. However, there was little apparent correlation between the expression of MMOs and k _{1, TCE} as the Cu²⁺ concentration increased. For example, the maximum TCE degradation rate occurred at a Cu²⁺ concentration of 5 μM, though there was no prominent expression of MMOs at this concentration, suggesting that enzymes other than MMOs played a part in TCE cometabolic degradation. Together, these observations suggested that changes in the diversity of the bacterial community induced by different Cu²⁺ concentrations resulted in different levels of MMOs and other enzymes involved in the cometabolic degradation of TCE.

Figure 4.

Relationship between first‐order kinetics constants ($k_{\mathrm{g},\mathrm{C}{\mathrm{H}}_4}$ and k _1,TCE) and relative expression of target genes (pmoA/16S and mmoX/16S).

3.3. Naphthalene transformation and the effect of Cu²⁺ supply on sMMO‐specific activity

As previously reported, sMMO activity is repressed at high Cu²⁺ concentrations in pure methanotrophs 13, 14, 15. In order to detect the activity of sMMO, a rapid colorimetric assay was performed at different Cu²⁺ concentrations. The results of color formation are shown in Supporting Information Fig. 3. The colored product was clearly visible, and the formation of colored diazo‐dye complexes was assessed by measuring the absorbance (A525) via spectrophotometry. The results were translated into naphthol concentrations using a standard curve (Supporting Information Fig. 4). Naphthol concentrations are shown in Supporting Information Table 4, and the changes in sMMO activity (expressed as nmol of naphthol formed per mg of cell protein per hour) relative to Cu²⁺ concentration are shown in Fig. 5. Compared with Methylomonas methanica strain 68‐1, which has a naphthol formation rate of 0–225 nmol/(mg_cell h) under Cu²⁺‐free conditions and no detectable sMMO activity in the presence of 1 μM CuSO₄ 43, the naphthol formation rate was higher in this study, ranging from 74.41 nmol/(mg_cell h) in 15 μM Cu²⁺ to 654.99 nmol/(mg_cell h) in 0.03 μM Cu²⁺. This suggested that sMMO produced by methanotrophs in MC retained higher activity levels at increased concentrations of Cu²⁺ 44. This finding reaffirmed that the bacteria mainly responsible for Cu²⁺ uptake in MC were likely heterotrophs rather than methanotrophs. Although sMMO activity was detected under various Cu²⁺ concentrations, it was substantially repressed at Cu²⁺ concentrations over 0.75 μM. The maximum activity of sMMO occurred at 0.03 μM Cu²⁺, which is consistent with the expression of the mmoX gene as well as TCE biodegradation. It should be noted that despite previous studies providing insights into the regulatory role of Cu²⁺ in pure cultures of methanotrophs, the regulatory role of Cu²⁺ in methanotrophic consortia is not known due to multiple interactions, including Cu²⁺‐mediated effects on non‐methanotrophs and the relationship between heterotrophic microorganisms and methanotrophic microorganisms. As such, it is still unclear how the sMMO and pMMO gene clusters are reciprocally regulated with respect to Cu²⁺. It has been speculated that the Cu²⁺‐specific‐binding compound in methanotrophs, methanobactin, may be the mechanism by which Cu²⁺ is sensed by these cells and may help coordinate the reciprocal regulation of sMMO and pMMO 13, 45.

Figure 5.

Naphthalene‐naphthol assay for sMMO activity.

3.4. Overall analysis of pyrosequencing

In order to determine the microorganisms involved in CH₄ uptake and TCE degradation, the microbial community structure in the samples at Cu²⁺ concentrations of 0.03 and 5 μM were investigated. Diversity indices and sequence information is shown in Table 1. The reads of the two samples were 10799 and 11964, respectively, and their coverage level reached 1.0000 and 0.999, respectively. This indicated that the profundity of sequencing was complete. Supporting Information Figure 5 illustrates the phylum level distributions of the bacterial Operational Taxonomic Units (OTUs) involved in the samples assayed at 0.03 μM (Supporting Information Fig. 5A) and 5 μM (Supporting Information Fig. 5B) Cu²⁺. It is apparent that most of the bacteria in both samples were grouped into Proteobacteria, Firmicutes, Actinobacteria, Bacteroidetes, and Verrucomicrobia. Similarly, based on PCR denaturing gradient gel electrophoresis (DGGE) fingerprinting or Illumina sequencing, prior research found that Proteobacteria and Actinobacteria were the dominant phyla in landfill cover soil 46. Our results indicated that Proteobacteria, Firmicutes, Actinobacteria, Bacteroidetes, and Verrucomicrobia at 0.03 μM Cu²⁺ accounted for 58.5, 28.7, 1.7, 10.0, and 1.0% of total bacterial sequences, respectively, whereas these phyla at 5 μM Cu²⁺ were 81.9, 14.5, 0.9, 2.4, and 0.2%, respectively. These results indicated that as Cu²⁺ increased the relative abundance of Proteobacteria also increased, and this was paralleled by a decrease in other phyla in the MC.

Table 1.

Diversity indices and sequence information from high‐throughput sequencing of experimental samples

Copper concentration (μM)	Reads	0.97
		OTUs	Ace	Chao 1	Coverage
0.03	10799	62	62	62	1.0000
5	11964	52	52	52	0.9999

3.5. Effect of Cu²⁺ on community structure involved in TCE degradation

Figure 6 shows the composition of the bacterial community in the two samples at the genus level. A significant difference in community structure was apparent between the two samples. The dominant microorganism was Methylocystaceae with a relative abundance of 36.1% and 75.4% at Cu²⁺ concentrations of 0.03 μM (Fig. 6A) and 5 μM (Fig. 6B), respectively, suggesting that higher Cu²⁺ concentrations significantly stimulated growth of methanotrophs. In addition, other microorganisms also occupied a certain proportion in the MC, including Lactococcus (13.1 and 6.3% at 0.03 and 5 μM, respectively), Bacillus (7.9 and 4.3% at 0.03 and 5 μM, respectively), Solibacillus (5.1 and 2.7% at 0.03 and 5 μM, respectively), Methylophilus (7.4 and 0.9% at 0.03 and 5 μM, respectively), and Taibaiella (5.2%, only at a Cu²⁺ concentration of 0.03 μM). It has been shown that these heterotrophs also degrade TCE using enzymes other than MMOs 47, 48, 49. Kaushik and Pranab isolated a novel strain belonging to the genus Bacillus than can utilize TCE as the sole carbon source 48. As such, based on the diversity and abundance of the community structure, it is apparent that many microorganisms play a role in TCE biodegradation in MC and that the dominant microorganisms and major enzymes involved in TCE biodegradation were altered at different Cu²⁺ concentrations.

Figure 6.

Genus‐level distributions of bacterial populations at 0.03 μM (A) and 5 μM (B) Cu²⁺.

Methanotrophic bacteria form close interactions with other heterotrophic bacteria 50, 51. In such consortia, methanotrophs form the basis of a methane‐driven food web, providing carbon compounds derived from substrates to other organisms 27. At different Cu²⁺ concentrations, van der Ha et al. 51 studied the selection of associated heterotrophs by methanotrophs. The results showed that different Cu²⁺ concentrations did not create major shifts in the methanotrophic communities, as a Methylomonas sp. was able to establish dominance at all Cu²⁺ concentrations and the associated heterotrophic communities showed continuous shifts, which is consistent with our current observations indicating that methanotrophic microorganisms are always dominant. It is proposed that methanotrophs select for certain heterotrophs, possibly fulfilling vital processes such as removal of toxic compounds or heavy metals 51.

4. Concluding remarks

For this study, the effects of Cu²⁺ on biodegradation and the microbiome of MC were investigated. An increase in Cu²⁺ concentrations from 0 to 15 μM altered the specific first‐order rate constant k _1,TCE, the expression levels of MMO genes (pmoA and mmoX), the specific activity of sMMO, and the microbial community composition of MC. Highly efficient TCE degradation (95%) and increased expression levels of MMOs were detected at a Cu²⁺ concentration of 0.03 μM. The sMMO activity of MC at Cu²⁺ concentrations of less than 15 μM was constantly present, which is inconsistent with prior studies of pure methanotrophs indicating that sMMO activity is depressed at high concentrations of Cu²⁺. These results indicated that the regulatory mechanism of Cu²⁺ in MC is distinct from that in pure cultures of methanotrophs. The results of MiSeq pyrosequencing showed that Cu²⁺ simulated the growth of methanotrophs and that Methylocystaceae was dominant bacteria.

Practical application

Methanotrophic consortia have been widely used in landfill cover for the reduction of greenhouse gases and have a strong application potential in the biodegradation of volatile organic compounds (VOCs). Notably, copper plays a key role in the regulation of the enzyme activities present in methanotrophic consortia that promote the biodegradation of contaminants. Therefore, this study evaluates the effects of copper on the regulation of methane monooxygenases (MMOs), trichloroethylene (TCE) degradation, and the community structure of methanotrophic consortia. The most effective copper concentration for highly effective degradation of TCE and expression of MMOs was determined. This research is expected to help in the development of a Cu²⁺‐mediated method for the treatment of VOC‐containing waste water or waste gases utilizing methanotrophic consortia.

The authors have declared no conflict of interest.

Supporting information

Supporting Information