Microbiome analysis of landfill leachate samples sourced from Powerstown landfill Co. Carlow

doi:10.21203/rs.3.rs-1507060/v1

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Microbiome analysis of landfill leachate samples sourced from Powerstown landfill Co. Carlow

https://doi.org/10.21203/rs.3.rs-1507060/v1

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Little is known about the microbial ecology of closed landfill, or how the diversity of these communities changes over time. This study focuses on the bacterial community diversity analysis of landfill leachate (LFL) sourced from Powerstown landfill Co. Carlow, Ireland. The LFL samples used in this study came from both closed and active landfill, and represent different phases of the landfill. HiSeq sequencing systems based 16S rRNA gene sequence analysis was used to investigate the bacterial communities of a total of six LFL samples; four samples from landfill cells (two from closed landfill cells and two from active landfill cells), and two samples sourced from a leachate tank (LT) collected over an interval of 2.5 years. Members of the phylum Proteobacteria (Pseudomonadota) were dominant in all landfill samples investigated (56 to 71%), followed by Bacteroidetes (Bacteroidota) (7 to 21%). The operating status of landfills (i.e. active / closed) and the age of landfill were found to affect the bacterial composition of LFL. Members of the class Epsilonproteobacteria and genus Sulfurimonas were dominant in LFL from active landfill, while the class Betaproteobacteria and genus Pusillimonas were dominant in closed landfill. Furthermore, it was observed that the composition of the microbial community in active landfill showed fluctuation by the progressing landfill age, while the bacterial composition of closed landfill tends to be stable as the year the landfill is closed.

Landfill leachate

Bacterial community diversity

16S rRNA sequencing

Proteobacteria (Pseudomonadota)

The disposal of municipal solid waste (MSW) is a global environmental problem being compounded by rapid population growth and the development of cities and industries. Landfilling is widely accepted as a method of treating MSW worldwide, largely due to its economic advantages. However, the production of landfill leachate (LFL), a wastewater rich in toxic organic and metal compounds and bacterial pathogens, by this process represents an environmental issue [1]. Many countries have gradually reduced the use of this waste disposal method. However, a large amount of waste is still disposed of in landfills every year, especially in some populous countries, for example in 2020, the USA landfilled 139.6 million tons of MSW [2], while 98 million tons of MSW were landfilled in China [3].

The organic fraction of MSW is degraded into CO₂, CH₄ and some low molecular weight compounds through a series of aerobic and anaerobic processes. The microbial communities which play a significant role in this degradation process include ammonia oxidizing bacteria, nitrite oxidizing bacteria, denitrifying bacteria, acidogens, and methanogens [4]. Due to the abundance of organic matter and substrate complexity in these sites, landfills are considered "microbial pools" [5], which are responsible for degrading the raw leachate and conversion of organic matter into less complex and less toxic compounds. However, the existence of bacterial pathogens in LFL may pose a threat to human and environmental health. Bioremediation of this waste stream may reduce the cost of treating leachate and the risk of secondary damage to the environment. Thus, the knowledge of the structure and composition of these microbial populations is necessary;

(1) to efficiently assess the public health risk,

(2) to optimize the operational parameters of LFL treating bioreactor systems.

Many studies have investigated the microbial composition of landfills and LFL. Boothe, Smith, Gattie, and Das [6] isolated and identified a series of both Gram-positive and -negative bacteria from LFL using plate counting technology and the Minitek microbial identification system. This study discovered a series of Gram-positive spp. including Bacillus megaterium, Bacillus pasteurii, Staphylococcus lentue, Staphylococcus delphini and other Gram-negative spp. that mainly belong to the genera Acinetobacter, Pseudomonas, Yersinia and Enterobacter. Similarly, bacteria including Bacillus spp., Staphylococcus spp., Lysinibacillus spp., Brevibacillus spp. and Clostridium spp. were isolated from LFL produced in Chandigarh, northern India by Krishnamurthi and Chakrabarti [7]. Song, Wang, Tang, and Lei [5] conducted a microbiological analysis of five landfill sites in China and found that Gammaproteobacteria, Firmicutes (Bacillota) and Bacteroidetes (Bacteroidota) were the most predominant bacteria among all landfill sites.

Although several investigations have been performed to characterize the microbial populations of LFL, a very limited number of studies have focused on the microbial communities involved in the humification of aged LFL from closed landfill sites, the changes in these communities over time or the pattern of bacterial community diversity through progressing landfill age. Therefore, the aims of this study were to; (1) Identify the dominant phyla/classes/genera in LFL, (2) Identify the changes in these microbial populations over time and (3) compare the microbial community composition of LFL between closed and active landfill.

2.1 Site description and leachate collection

LFL used in this study was sourced from Powerstown landfill, Co. Carlow, Ireland (52°45’58.46’’ N, 6°57’20.13’’ W). The landfill is located 8 km south-east of Carlow Town in a rural setting and has been operational since 1975. The site consists of three different phases; phase 1 (P1) which operated from 1975–1990, phase 2 (P2) operated from 1991–2006, which covers approximately 5.7 hectares and has 13 engineered cells, and phase 3 (P3) opened in 2006 and closed before the end of 2018, including four lined cells, a surface water settlement pond, a leachate tank and a green waste composting area. Leachate collection systems were in operation in both active and closed landfills. Four samples were collected from landfill cells 11, 13, 16 and 18, and two samples were collected from the same leachate tank. Cell 11 (C11) and cell 13 (C13) belong to P2 of Powerstown (1991–2006), while C11 was closed 2 years before C13; Cell 16 (C16) and cell 18 (C18) belong to P3 (2006–2019); Sample LT and LT1 correspond to the leachate tank but in different sampling time, the information of the samples used in this study is listed in Table 1, all samples were stored at 4°C prior to analysis.

Table 1

Sample identification and details
Sample name	Source location	Phase	Operating status	Years operated	Years closed	Date of sampling	Comment
C11	cell 11	2	Closed	13 years	11 years	March 2018	-
C13	cell 13	2	Closed	15 years	9 years	March 2018	-
C16	cell 16	3	Active	9 years	-	March 2018	-
C18	cell 18	3	Active	9 years	-	March 2018	Newly fulfilled before sampled
LT	Leachate tank	3	Active	9 years	-	November 2015	-
LT1	Leachate tank	3	Active	11.5 years	-	March 2018	-

2.2 DNA extraction and 16S rRNA Gene Library Preparation

Total genomic DNA was extracted from all samples using the CTAB/SDS method [8]. Extracted DNA was diluted to 1ng/µL using sterile water. 16S rRNA genes of distinct regions (16SV3-V4) were amplified used a specific primer with the barcode (Online Resource Table 1). All PCR reactions were carried out with Phusion® High-Fidelity PCR Master Mix (New England Biolabs). Equal volumes of 1X loading buffer (containing SYB green) was mixed with PCR products and ran on 2% agarose gel for detection. PCR products between 400-450bp were chosen for further experiments. PCR products were mixed in equidensity ratios. PCR products were purified with Qiagen Gel Extraction Kit (Qiagen, Germany). Libraries were generated with NEBNext® Ultra™ DNA Library Prep Kit for Illumina, following manufacturer's recommendations and index codes were added. Library quality was assessed on the Qubit 2.0 Fluorometer (Thermo Scientific) and Agilent Bioanalyzer 2100 system. The library was sequenced on an Illumina platform and 250 bp paired-end reads were generated.

2.3 Sequence Analysis

Paired-end reads were assigned to samples based on their unique barcode and truncated by cutting off the barcode and primer sequence. Paired-end reads were merged using FLASH [9]. Quality filtering of raw tags was performed under specific filtering conditions to obtain the high-quality clean tags [10] according to the Qiime quality-controlled process [11]. Sequences analysis was performed by UPARSE [12] software using all the effective tags. Sequences with ≥ 97% similarity were assigned to the same operational taxonomic units (OTUs). Representative sequence for each OTU was screened for further annotation. For each representative sequence, the RDP Database [13] was used to annotate taxonomic information. A representative sequence of each OUT was aligned with pyNAST [14] and filtered to remove uninformative bases. The top 100 genera were selected and evolutionary trees were drawn using the aligned represent sequences. Multiple sequence alignments were conducted using the MUSCLE software [15] (Version 3.8.31), and used for community composition analyses. Rarefaction curves and multiple diversity metrics including Abundance-based Coverage Estimator (ACE), Chao1, the Good’s coverage and the Shannon index were used to estimate the bacterial richness and diversity in the samples. Principal Coordinate Analysis (PCoA) was performed to get principal coordinates and visualize from complex, multidimensional data. A distance matrix of unweighted UniFrac [16] among samples obtained before was transformed to a new set of orthogonal axes, by which the maximum variation factor is demonstrated by first principal coordinate, and the second maximum one by the second principal coordinate, and so on. A core microbiome was computed for all samples within QIIME. A mapping file is included as Online Resource Table 1, and the commands used to produce the final BIOM file is displayed in appendix 1.

3.1 Estimation of richness and diversity

After quality filtering and removal of chimeric sequences, a total of 203140 high-quality reads were obtained: 36405 for C11, 31108 for C13, 35767 for C16, 32924 for C18, 34251 for LT and 32685 for LT1, with an average length of 419 bp (Online Resource Table 2). The sequences clustered into 2581–5739 OTUs per sample when applying a 97% similarity threshold for inclusion.

The rarefaction analysis of the six leachate samples at an OTU cut-off value of 97% indicated that a full census had not been achieved for any of the samples (appendix Fig. 4), all the curves were far from reaching a plateau, suggesting that the curves largely underestimate the numbers of different types of bacteria in each sample, which did not allow for calculating the total species richness, the description of biological diversity only based on all the identified OTUs.

A series of coverage estimators were computed, including ACE, Chao1, the Good’s coverage, and the Shannon index (Table 2). Good’s estimator indicated nearly complete coverage. The sequence composition analyses revealed that the bacterial diversity estimates for the LFL were much greater than previously published descriptions of LFL diversities. The average number of OTUs for all samples was 3680, which was much greater than that found in previous studies. For example, Song, Wang, Tang, and Lei [5] investigated the bacterial communities of ten LFL samples from five landfill sites in China and found 709–1599 OTUs per sample. The community richness (Chao 1 and ACE) showed the highest for C13 and the lowest for LT1, which correlates with the number of observed OTUs in the samples; the community diversity (Shannon’s H index which considers the evenness of OTU distribution) showed the highest for C13 and the lowest for C16.

Table 2

Coverage estimates for the bacterial LFL communities
Sample	OTU richness	ACE*	Chao1	Good’s coverage	Shannon
C11	3545	13553	13226	0.983	4.225
C13	5739	28221	24135	0.986	4.647
C16	2960	13038	11812	0.986	3.251
C18	4234	16480	15116	0.985	4.019
LT	3019	12953	11547	0.985	3.481
LT1	2581	12530	10513	0.985	3.263
*ACE abundance-based coverage estimator

3.2 Diversity patterns and impact factors

The community composition of the six samples were compared using a PCoA analysis based on unweighted UniFrac distance between samples, the top two variation factors are shown in Fig. 1. In combination, the two PCoA axes (PC1 and PC2) explained 45.9% of the variance in microbial community abundance between samples. The result indicated that the status of landfill (closed landfill / active landfill) might affect the microbial community composition, LT1 / C16 and LT (which belong to active landfill) formed a distinct cluster, C11 and C13 (closed landfill) showed close distance with each other, while C18 (active landfill but being newly filled) appeared between those two clusters. The years of landfill operated / closed, also contributed significant effort for the similarity of landfill bacterial community, which caused the separation between C11and C13 (LT1 and LT) on axis PC2.

3.3 Microbial community structure

The comparative analysis using the RDP (Ribosomal Database Project) database showed that only 14 to 105 of the sequence tags (0.04 to 0.29%) were affiliated with the domain Archaea. Due to their extremely low abundance, these sequences were excluded from the community structure analysis.

For all six samples (four from landfill cells and one from LFL tank, the most dominant phylum was Proteobacteria (Pseudomonadota) (56–71%, Fig. 2A and Table 3). This was also reflected in the evolutionary trees of the top 100 genera, where the most abundant genera belonged to the phylum Proteobacteria (Pseudomonadota) (appendix Fig. 5). The next most abundant phyla were Bacteroidetes (Bacteroidota) (7 to 21%), Actinobacteria (Actinomycetota) (4 to 9%), Firmicutes (Bacillota) (5 to 8%) and Balneolaeota (Balneolota) (2 to 6%), as shown in Fig. 2A. The remaining phyla that featured in the top 10 abundances for each sample included Spirochaetes (Spirochaetota), Cyanobacteria, Tenericutes, Euryarchaeota, Gemmatimonadetes, Verrucomicrobia (Verrucomicrobiota), Acidobacteria (Acidobacteriota), Synergistetes (Synergistota) and Chloroflexi (Chloroflexota), however, it should be noted that their individual contribution only varied between 0.15 to 2.07% of the bacterial sequence set of each sample (Table 3). The microbial community composition of the top 10 proportions of each sample from superkingdom (L1) to genus (L6) is shown in Online Resource Table 5–8.

Table 3

HiSeq sequencing results (Top 10 Hits Distribution)
Sample	Acido.	Actino.	Bact.	Baln.	Cyano.	Chlo.	Eury.	Firm.	Gemma.	Proteo.	Spir.	Syne.	Tener.	Verr.
C11	-	2689	2527	2304	248	-	105	1902	96	25268	752	-	130	-
C13	170	1927	2312	1093	344	-	-	1655	396	22092	335	-	-	253
C16	-	2073	7216	784	54	-	-	1855	235	22745	230	155	70	-
C18	-	3113	2384	2111	319	93	-	2436	72	21554	402	78	-	-
LT	-	1251	8058	1005	391	105	-	2636	209	19319	638	124	-	-
LT1	-	2210	6984	871	121	66	-	1537	186	19763	472	100	-	-
AcidoAcidobacteria (Acidobacteriota), Actino Actinobacteria (Actinomycetota), Bact Bacteroidetes (Bacteroidota), Baln Balneolaeota (Balneolota), Cyano Cyanobacteria, Chlo Chloroflexi (Chloroflexota),

Eury Euryarchaeota, Firm Firmicutes (Bacillota), Gemma Gemmatimonadetes (Gemmatimonadota), Proteo Proteobacteria (Pseudomonadota), Spir Spirochaetes (Spirochaetota), Syne Synergistetes (Synergistota), Tener Tenericutes, Verr Verrucomicrobia (Verrucomicrobiota)

The result of OTUs clustering, where both the common and unique information for different samples were analysed is represented in a flower diagram (Fig. 3). This analysis indicated that all samples shared 292 core OTUs. The majority of these OTUs belong to Betaproteobacteria, Gammaproteobacteria, Alphaproteobacteria and Bacteroidetes (Bacteroidota), reflecting the predominance of the Proteobacteria (Pseudomonadota) and Bacteroidetes (Bacteroidota), also observed during whole community analysis (Table 3). In addition, the number of unique OTUs in C11, C13, C16, C18, LT and LT1 were 63, 140, 46, 62, 46 and 32, respectively (Fig. 3).

Each petal in the flower diagram represents a sample. The core number in the centre is for the number of OTUs present in all samples, while the number in the petals for the unique OTUs only showing in each sample.

Active landfill (C16, C18&LT) and the closed landfill (C11&C13)

Spp. from the phylum Proteobacteria (Pseudomonadota) occupied a significant proportion (up to 71%) of the communities present in both C11 and C13 (landfill cell 11 and landfill cell 13, both closed landfill sites), followed by Bacteroidetes (Bacteroidota) (up to 7%), Actinobacteria (Actinomycetota) (up to 7%), Firmicutes (Bacillota) (up to 5%) and Balneolaeota (Balneolota) (up to 6%) (Fig. 2A and Online Resource Table 4). Samples C16, C18 and LT were sourced from active landfill sites, while C18 was newly filled before sampled, sample LT came from the leachate tank that collected leachate from all active landfill sites. However, the new filled waste caused the microbial composition of C18 to be very different from C16 and LT. Interestingly, the most predominant phyla in C18 were Proteobacteria (Pseudomonadota), Actinobacteria (Actinomycetota), Bacteroidetes (Bacteroidota), Firmicutes (Bacillota) and Balneolaeota (Balneolota), and the most predominant classes were Betaproteobacteria and Gammaproteobacteria, Actinobacteria and Alphaproteobacteria (Table 3), which showed similar bacterial composition with C11 and C13. For samples C16 and LT, the top five phyla were similar to those found in C11 / C13, however, the ratio of each phylum in these samples (C16 and LT) showed a significant difference. The proportion of phylum Proteobacteria (Pseudomonadota) decreased from 69% (C11) / 71% (C13) to 63% (C16) / 56% (LT) mainly caused by the dramatic decreasing of class Betaproteobacteria, Gammaproteobacteria and Alphaproteobacteria, the representation of class Betaproteobacteria dropped from the highest 33.61% (C13) to the lowest 9.4% (C16), as shown in Online Resource Table 8. Meanwhile, it is worth noting that the proportion of class Epsilonproteobacteria in the samples of active landfill sites (C16 and LT) has increased significantly compared with that of closed landfill sites (C11 and C13), class Epsilonproteobacteria has become the most dominant class in C16 and LT. Due to this behaviour, the total Proteobacteria (Pseudomonadota) in samples C16 and LT could be able to remain the most dominant phylum. Contrary to the trend of phylum Proteobacteria (Pseudomonadota), the proportion of phylum Bacteroidetes (Bacteroidota) has increased in active landfill compared to closed landfill, from the lowest 7% (C11) to the highest 23% (LT), members in class Cytophagia contributed main effort for this result, increased from less than 1% in samples C11 and C13 to around 15% in samples C16 and LT (Online Resource Table 5). Genera Pusillimonas and Sulfurimonas, which both are affiliated with Proteobacteria (Pseudomonadota), were predominant at samples in active and closed landfills, respectively; Pusillimonas representing 20% and 21% of the total genera for C11 and C13, Sulfurimonas accounting for 40% and 36% of the total genera for C16 and LT1 (Online Resource Table 8). The second dominant genus for C11 (8%) and C13 (7%) was Pseudomonas, while the genus Spirosoma, which belongs to phylum Bacteroidetes (Bacteroidota) was the second dominant genus for C16 (14%) and LT (16%) (Online Resource Table 8). Interestingly, Spirosoma were not present in the closed LFL samples, Pseudohongiella accounted for 6% and 4% of the genera at C11 and C13 but were not present in active landfill samples (Fig. 2C).

C11 and C13

Both C11 and C13 were sampled from closed aged of landfill cells, with C11 closed two years before C13. Subtle differences between the microbial composition of both samples were observed. Members of the phylum Proteobacteria (Pseudomonadota) occupied a lower proportion of the community present in sample C11 compared to C13, which is caused by the combined effect of an increase of Betaproteobacteria and Alphaproteobacteria and a decrease of Gammaproteobacteria. In contrast, the proportion of Balneolaeota (Balneolota) increased from 3% (C13) to 7% (C11), and the ratio of Spirochaetes (Spirochaetota) increased from 1% (C13) to 2% (C11), other phyla did not show obvious fluctuation between two samples (Online Resource Table 4). At genus level, Pseudomonas (Gammaproteobacteria), Candidatus Cyclonatronum (Balneolaeota (Balneolota)), Pseudohongiella (Gammaproteobacteria) and Leucobacter (Actinobacteria) displayed higher percentage in C11 compared with C13 (Fig. 2C).

LT and LT1

LT and LT1 were collected from the same leachate tank, however, LT1 was sampled 2.5 years before LT. The microbial composition of the two samples presented a significant difference. Proteobacteria (Pseudomonadota) experienced significant growth from 56% (LT) to 60% (LT1), mainly contributed by the increase of Epsilonproteobacteria and Betaproteobacteria. Similarly, Actinobacteria (Actinomycetota) increased from 3% (LT) to 6% (LT1) (Online Resource Tables 4 and 5). In contrast, Bacteroidetes (Bacteroidota) decreased from 23% (LT) to 21% (LT1), which was principally affected by the decrease of Cytophagia and Bacteroidia; with Firmicutes (Bacillota) decreasing from 8% (LT) to 5% (LT1) (Online Resource Tables 4 and 5). At the genus level, Sulfurimonas, which belongs to phylum Bacteroidetes (Bacteroidota), occupied the highest proportion of the bacterial communities present in both samples (36% (LT) and 39% (LT1)). Pusillimonas, from the phylum Proteobacteria (Pseudomonadota), remained constant at 5% (LT) and 6% (LT1). While Spirosoma, from the phylum Bacteroidetes (Bacteroidota), was the second most abundant genus in both samples, decreased with progressing landfill age, from 16% in LT to 14% in LT1 (Online Resource Table 8).

The OTUs and diversity index values in this study appeared much higher than those determined in previous studies, indicating that the bacterial diversity in landfills is greater than previously thought. The phyla, Bacteroidetes (Bacteroidota), Actinobacteria (Actinomycetota), Firmicutes (Bacillota), Balneolaeota (Balneolota), Cyanobacteria, Euryarchaeota, Gemmatimonadetes, Synergistetes (Synergistota), Chloroflexi (Chloroflexota) and Spirochaetes (Spirochaetota) have been detected in previous reports of landfill or LFL as well as in this study [17–21]. The most dominant phylum found in this study was Proteobacteria (Pseudomonadota), which is consistent with the results of the study conducted by Stamps et al. [22] from 19 landfill sites in the USA. However, Köchling, Sanz, Gavazza, and Florencio [23] analyzed the microbial community structure of the leachate from a Brazilian landfill and found that the dominant phyla were Firmicute, Song, Wang, Tang, and Lei [5] found out that the phyla of Fusobacteria and Tenericutes were in a dominant position from landfills in China. This suggests the microbial community structure in landfills is strongly influenced by environmental conditions and their substrate specificity, which explains the difference in the results shown in different studies.

Surprisingly the bacterial community observed in C18, which was sampled from the leachate from the active landfill which had been newly filled, showed a high similarity with the ageing landfill, and a significant difference from the other samples from the active landfill. As shown from the PCoA (Fig. 1), its composition was between the aged landfill group and the active landfill group, indicating that the newly landfilled waste greatly affected its bacterial composition. Thus, it is not considered into an active landfill group for discussion.

Although phyla Proteobacteria (Pseudomonadota) and Bacteroidetes (Bacteroidota) were dominant in both groups, the proportion of Proteobacteria (Pseudomonadota) was higher in leachate recovered from closed landfill (70 ± 0.91%) compared to active ones (60 ± 5.1%). In contrast, the phylum of Bacteroidetes (Bacteroidota) appeared at a higher ratio in the leachate of active landfill (21 ± 1.85%) than in closed landfill (7 ± 0.26%). Two groups also showed completely different compositions from class, order, family and genus level, the most dominant in active landfill group were Betaproteobacteria, Burkholderiales, Alcaligenaceae and Pusillimonas in these levels, respectively, while for the closed landfill group they were Epsilonproteobacteria, Campylobacterales, Helicobacteraceae and Sulfurimonas. The above findings suggest significant variance for bacterial compositions between leachate from active and closed landfill. The major genera identified in leachate from closed landfill, including Pusillimonas, Pseudomonas, Pseudohongiella and Candidatus Cyclonatronum, accounted for around 40% of the entire bacterial population. Genus Pusillimonas which appeared to have a major role in the stabilization process of the aged LFL, represented more than 20% in taxa composition in both samples. This result shows similarity with the microbiome analysis of an old landfill (closed for two years) carried out by Remmas, Roukouni, and Ntougias [24], who found a large fraction of Pusillimonas-like bacteria which represented c. 34% of the total reads in samples. Pusillimonas strains were mentioned as “bacterial strains that had an extraordinary degradative ability” [25]. Lladó, Gràcia, Solanas, and Viñas [26] also indicated Pusillimonas strains were involved in the bioremediation of an aged creosote-polluted soil sourced from Barcelona. Members of Pseudomonas are one of the typical cellulolytic microorganisms and are well-known as a degrader of recalcitrant organic compounds [27]. For example, some Pseudomonas spp. are specialized in the biotransformation of hexavalent chromium [28]. Pseudohongiella and Candidatus Cyclonatronum increase with the increased years of landfill closure, unfortunately, they are novel bacteria that have just been isolated in recent years and their functions are unclear.

It is worth noting that the major microbiota identified in the aged landfill were mostly strict aerobes, including Pusillimonas, Pusillimonas and Pseudohongiella spp., indicating that the landfill has ended the methanogenic stage and air has begun to penetrate through the landfill surface (phases VI to VIII of the landfilling process-as reviewed by [29]). The time required for landfills to deplete all organic matter ranges from a few centuries (for an unsaturated, uncovered landfill) to more than 500,000 years for a landfill fully saturated with water (that is, waste placed below the groundwater level) [30]. The majority of the bacterial identified in these two samples were related to microbiota isolated from soil, a fact that also indicates the completion of the humification process. At this stage, humic substrates appear to be the major organic compounds identified in the LFL [30]. Qi, Yue, and Nie [31] reported that humic compounds in mature LFL are most similar in nature to humus-derived substances than to waste-derived humic acids, explaining the dominance of bacteria commonly found in terrestrial ecosystems.

For the active landfill group, the most abundant genera were Sulfurimonas, Spirosoma, and Pusillimonas, in which Sulfurimonas occupied almost 40% of the abundance in samples. The Sulfurimonas detected in this study is in accordance with a previous study of leachate sourced from landfills in China, which stated that Sulfurimonas spp. contribute significantly to the global sulfur cycle [5]. Han and Perner [32] also indicated that Sulfurimonas spp. are capable of reducing nitrate, oxidizing both sulfur and hydrogen, and containing group IV hydrogenases. As a sulfur-oxidizing Epsilonproteobacterium, studies have found that Sulfurimonas spp. use a wide variety of electron donors for growth including sulfide, sulfur, thiosulfate, and sulfite. Sulfur oxidizing Sulfurimonas spp. are widespread in sediments, hydrothermal vent fields, aquifers, and subsurface environments such as oil reservoirs where they play an important role in the sulfur cycle [33]. Spirosoma ssp. exist in some common habitats, such as fresh water and soil [34, 35], and have also been isolated from some extreme environments, such as high Arctic permafrost soil [36], and Zn- and Cd-accumulating Salix caprea (pussy willow) [37]. The roles of Spirosoma spp. are still being explored with a recent paper reporting Spirosoma sordidisoli capable of degrading the common herbicide propanil [38].

It can be seen from the analysis results of bacterial composition between LT and LT1, that the bacterial composition showed fluctuation by the progressing landfill age in phyla, class, order, family and genus levels. This phenomenon is consistent with the research of Köchling, Sanz, Gavazza, and Florencio [23] and Song, Wang, Tang, and Lei [5], however, as the microbial composition of LFL is sensitive to several factors including environment and nutrient substrate composition [30], the trend of bacterial communities in LFL with landfill age obtained from various studies are not comparable. In the closed landfill group, the bacterial communities only showed slight variation for samples with 2 years’ degraded time difference, it could be speculated that the landfill process had reached in the same phase in these two landfill cells and the bacterial composition has stabilized. It should be noted that unfortunately, due to very limited number of studies on the microbial communities involved in the humification of aged LFL from closed landfill sites, the changes in these communities over time or the pattern of bacterial community diversity through progressing landfill age, especially in Ireland, there was little to compare this study to.

This study investigates the bacterial communities present in a range of LFL samples and identifies the dominant phyla/classes/genera therein. In particular, this study focuses on the bacterial composition of LFL from closed aged landfill, which enriches the current research deficiencies in this field and provides theoretical support for leachate treatment. The comparison of bacterial communities in LFL between active and closed landfills indicates that the operating status of landfill contributes significantly in terms of leachate microbiomes composition. Furthermore, the changing patterns of microbial components in LFL over time are displayed, with the microbial composition of LFL from active landfill varying by the growing landfill age, which tends to stabilize with time in LFL from closed landfill. Specifically, the proportion of Epsilonproteobacteria, Betaproteobacteria, Actinobacteria (Actinomycetota), Bacteroidetes (Bacteroidota) and Firmicutes (Bacillota) varied in samples LT and LT1 (from active landfill sites but with 2.5 years age difference). However, the microbial composition of samples from closed LFL (C11 and C13) did not show a significant difference even with 2 years age difference.

In addition, this study confirms that the microbial communities present in LFL contain a large number of novel strains making LFL an excellent location to isolate new strains. Importantly this diverse specialized bacterial community is capable of degrading recalcitrant organic compounds and resisting high HM concentrations accumulated in these environments, as described above, i.e., Pusillimonas sp., Pseudomonas sp. and Spirosoma sp.. Similarly, many studies have been reported that isolates from LFL showed resistant to HMs or shown bioremediation ability to HMs. For example, in a recent study, Oziegbe, Oluduro, Oziegbe, Ahuekwe, and Olorunsola [39] found isolates from landfill (Pseudomonas aeruginosa, Klebsiella edwardsii and Enterobacter cloacae) with tolerance to Cd and Pb, and Pseudomonas aeruginosa exhibiting the bioremediation ability with 58 and 33 remediation percentage in 50 mg Cd L^− 1 and 300 mg Pb L^− 1; Imron, Kurniawan, and Soegianto [40] isolated Pseudomonas aeruginosa from LFL in Indonesia, this starin was capable of removing Hg with a removal percentage of 99.7 ± 0.18% in a 10‰ saline medium; Moris, Garcia-Cabellos, Enright, Ryan, and Enright [41] also screened out Bruvundimonas diminutas, Brevundimonas naejangsanensis, Brevibacterium iodinum and Bacillus cereus from LFL samples according to their tolerance to Cu, Ni, Fe, Cd and As. These spp. could represent a great biological source for bioremediation, and the screening of microbial spp. for efficient pollutant degrading is a critical step in the development of this systems.

Author contributions

All authors contributed to the study conception and design. Huili He and Sinead Morris performed material preparation, data collection and analysis, Guiomar Garcia-Cabellos and Ann-Marie Enright contributed reagents/materials/analysis tools to the work. The first draft of the manuscript was written by Huili He and all authors commented on previous versions of the manuscript. All authors read and approved the final manuscript.

Acknowledgements

This study was carried out at EnviroCORE, Department of Science and Health, Institute of Technology Carlow, Ireland. We acknowledge support from the Institute of Technology Carlow President’s Research Fellowship Programme fund.

Declarations

The research leading to these results received funding from Institute of Technology Carlow President’s Research Fellowship.

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Microbiome analysis of landfill leachate samples sourced from Powerstown landfill Co. Carlow

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Abstract

Figures

1. Introduction

2. Materials And Methods

2.1 Site description and leachate collection

2.2 DNA extraction and 16S rRNA Gene Library Preparation

2.3 Sequence Analysis

3. Results

3.1 Estimation of richness and diversity

3.2 Diversity patterns and impact factors

3.3 Microbial community structure

4. Discussion

5. Conclusion

Declarations

References

Supplementary Files

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