Journal of Soils and Sediments https://doi.org/10.1007/s11368-018-2090-4
SEDIMENTS, SEC 2 • PHYSICAL AND BIOGEOCHEMICAL PROCESSES • RESEARCH ARTICLE
Differentiation of nitrogen and microbial community in the littoral and limnetic sediments of a large shallow eutrophic lake (Chaohu Lake, China) Weidong Wang 1 & Weiyue Liu 1,2 & Di Wu 1 & Xiaoxia Wang 1 & Guibing Zhu 1,2 Received: 21 April 2018 / Accepted: 12 July 2018 # Springer-Verlag GmbH Germany, part of Springer Nature 2018
Abstract Purpose Nitrogen (N) is one of the major elements causing eutrophication in freshwater lakes, and the N cycle is mainly driven by microorganisms. Lake littoral zones are found to be Bhotspots^ for N removal from both the basin and receiving waters. However, the environmental factors that drive the distribution of microorganisms are diverse and unclear. Here, we examined the differentiation of nitrogen and microbial community between the littoral and limnetic sediments to explore their interactions. Materials and methods Sediment samples were collected in the littoral and limnetic zones of Chaohu Lake in winter (ca. 7 °C) and autumn (ca. 22 °C). Abundances of the bacterial and archaeal genes amoA (ammoxidation), nirS and nirK (denitrification), hzsB (anaerobic ammonium oxidation; anammox), and nrfA (dissimilatory nitrate reduction to ammonium; DNRA) were measured via quantitative real-time polymerase chain reaction (qPCR). Clone libraries were constructed for further phylogenetic analysis to study the community composition. Results and discussion We observed significant higher concentration values in terms of sedimentary NH4+-N and NO3−-N in the limnetic zone than littoral zone (p < 0.05; n = 12). In general, abundance values of the above six genes in the littoral zone were all higher than those in the limnetic zone, while higher in winter (7 °C) than in autumn (22 °C) (p < 0.05; n = 6). The spatial heterogeneity had the most significant effect on the distribution of ammonia-oxidizing archaea (AOA) and anammox bacteria abundance. Both temporal (temperature) and spatial heterogeneity affected the abundance of ammonia-oxidizing bacteria (AOB). The variation in the abundance of denitrifying bacteria and DNRA bacteria mainly reflected the temporal (temperature) heterogeneity. Conclusions The six N-cycle-related microorganisms were affected by different environmental factors and presented different distribution patterns. The lower nitrogen content and the higher microbial abundance and diversity showed that the littoral zone was the Bhotspot^ of N-cycling-related microorganisms in a large, eutrophic, and turbid lake. It is suggested that increasing the area and restoring the ecological function of the littoral zone was effective and significant in eutrophic lake management. Keywords Abundance . Eutrophic freshwater lake . Littoral zone . Microbial N-cycle . Sediment
1 Introduction Responsible editor: Shiming Ding Electronic supplementary material The online version of this article (https://doi.org/10.1007/s11368-018-2090-4) contains supplementary material, which is available to authorized users. * Guibing Zhu
[email protected] 1
Key Laboratory of Drinking Water Science and Technology, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing 100085, China
2
University of Chinese Academy of Sciences, Beijing 100049, China
Freshwater eutrophication is a widespread global environmental problem and a consequence of intense human activities. In the Yangtze River Basin in China, an investigation of 138 lakes larger than 10 km2 showed that 85.0% of the surveyed lakes were eutrophic (Yang et al. 2010). Eutrophication and algal blooms (Microcystis spp.) have been occurring since the 1980s, with subsequent risks to human health due to the toxins produced by cyanobacteria (Chen et al. 2018). Nitrogen (N) and phosphorus (P) supplies contribute to freshwater eutrophication and are the result of a rapidly growing agriculture and intensified urban activities (Ding et al. 2018; Xing et al. 2018).
J Soils Sediments
The nitrogen cycle (N cycle) and nitrogen removal from waterbodies involve different reductive or oxidative reactions, mainly driven by microorganisms, and are of significant importance in controlling eutrophication (Hayatsu et al. 2008). Sediments are the main N sinks in lacustrine systems and can function as a secondary source of nutrients for eutrophication processes under some conditions (Ali et al. 1988). Thus, lake sediments can provide important historical information regarding nutrient levels, ecosystem evolution, and human activities (Haworth et al. 1984). Lake littoral zones, the transitional boundary zones between terrestrial and aquatic ecosystems, play important roles in intercepting pollutants of aquatic systems (Robert and Henri 1997). Wind-driven hydrodynamic forces, water current, and human activities result in the intensity of upwellings and drive water movements in the shallow parts of lakes, causing frequent material exchange between the sediments and the overlying water in the littoral zone (Buzzelli 1998; Alvarez-Cobelas et al. 2005). On the contrary, the limnetic zone sediments are located at relatively deeper levels and therefore subjected to less disturbance; they also showed lower N-cycling rates (Bruesewitz et al. 2012). However, our knowledge about N-cycling microorganisms in these two distinct zones, with significant physicochemical heterogeneity, is still limited, making comparative studies in this present field necessary. Ammonia oxidation, as the first and rate-limiting step of subsequent N transformations, is mainly performed by ammonia-oxidizing archaea (AOA) and bacteria (AOB) (Purkhold et al. 2000; Schleper and Nicol 2010). Venter et al. (2004) found that the ammonia monooxygenase (AMO) gene in the archaea drives ammoxidation. At present, we know that AOA and AOB coexist in most habitats, such as soils (Leininger et al. 2006), oceans (Francis et al. 2005), reactors (Kasuga et al. 2010), lakes (Wang et al. 2012; Zhou et al. 2015), and wetlands (Sims et al. 2012). However, there are dramatic differences in the ecological niche distribution of these two active ammonia-oxidizing microorganisms under different environmental conditions. For example, AOA are favorably adapted to low-ammonium conditions, while AOB can adapt to environments with high oxygen contents (Martens-Habbena et al. 2009; Prosser and Nicol 2012; PettRidge et al. 2013). The physicochemical heterogeneity of the littoral and limnetic zones is considerable, and the differences in abundance and community composition of AOA and AOB in these two zones are well known. Denitrification is the stepwise reduction of nitrate (NO3−) to dinitrogen (N2) and represents an important step in removing nitrogen from aquatic systems. Denitrifying bacteria are heterotrophic microorganisms and can use N oxides as terminal electron acceptors for cellular bioenergetic processes under anaerobic, microaerophilic, and, occasionally, aerobic conditions (Zumft 1997). The differences between the littoral and limnetic zones in terms of denitrifying bacterial
distributions are not well known. Recently, an alternative pathway of N removal, anaerobic ammonium oxidation (anammox), has been discovered. Anammox bacteria are ubiquitously distributed at redox transition zones in marine (Kuypers et al. 2003), freshwater (Schubert et al. 2006), and terrestrial ecosystems (Zhu et al. 2011). The land-freshwater interface represents a biogeochemical hotspot for anammox (Zhu et al. 2013). However, the key factors determining the distribution of different anammox bacteria are still not well known. Dissimilatory nitrate reduction to ammonium (DNRA) is the key process related to nitrate reduction, but without gas formation. Although it does not lead to N losses in aquatic systems, it plays an important role in balancing the N cycle. The percentage of nitrate consumed by DNRA (in terms of total nitrate consumption) differs greatly between ecosystems, with the lowest reported proportion of 2–4% (MA, North River) and the highest of 90–92% (Denmark, estuary sediment) (Megonigal et al. 2005). However, studies on DNRA in the littoral sediment zone of freshwater lakes are scarce. In this context, based on the amoA, nirS, nirK, hzsB, and nrfA genes involved in the abovementioned processes, we studied their distributions and the driving environmental factors. For this purpose, sediment samples were collected in the littoral and limnetic zones of Chaohu Lake, which presents a significant spatial heterogeneity during winter and autumn. Related functional genes were measured via quantitative real-time polymerase chain reaction (qPCR) and clone libraries were constructed. Abundances and community composition were further compared by redundancy analysis and phylogenetic analysis.
2 Materials and methods 2.1 Site description Chaohu Lake (117° 16′ 54″–117° 51′ 46″ E, 31° 25′ 28″–31° 43′ 28″ N) covering about 760 km2, located in Anhui Province of eastern China, is the fifth largest freshwater lake in China. The mean water depth is approximately 3 m. The catchment is subjected to subtropical monsoon climate with an annual average temperature of 15–16 °C and precipitation of 1100 mm. Chaohu Lake was typically dominated with algae during algal-bloom period rather than aquatic macrophytes. The western part of the lake is hypereutrophic, whereas the eastern part is eutrophic. Nanfei River, Paihe River, Hangbu River, and Baishitian River, which all flow into the western lake, provide more than 90% of runoff volume to the catchment, and also bring most of the pollutants (Yin 2011). Since the construction of the Yuxi Dam, the annual exchange of water between Chaohu Lake via Yuxi River and the Yangtze River has decreased from 13.6 × 108 m3 to 1.6 × 108 m3, which
J Soils Sediments
represents 45 and 5% of annual runoff from the whole catchment, respectively (Zan et al. 2012). In the past several decades, Chaohu Lake suffered from serious eutrophication resulting from increased anthropogenic nutrient input mainly from Nanfei River which discharges untreated industrial and municipal wastewater from Hefei City (capital of Anhui Province) located at the western part of Chaohu Lake. Water with a considerable quantity of pollutants flows from the western lake towards the eastern lake and finally settles in the sediments under a ca. 0.02–0.10 m s−1 wind-induced drift currents (Chen and Liu 2014). However, the eastern lake is the only drinking water intake for the Chaohu City at present, so it is urgent to study this area in order to improve the water quality.
2.2 Sediment sampling The average sediment depths of western and eastern Chaohu Lake are ca. 40 and 30 cm respectively. Surface sediment (upper 0–10 cm) samples were collected from the littoral zone
Fig. 1 Locations of Lake Chaohu and sampling sites. Sitemap showing Lake Chaohu’s drainage area and its surrounding major rivers and cities. Percentages indicate the proportion of runoff to the total. The arrow indicates the direction of water flow. Red circles show sampling
(non-macrophyte-covered near lakeshore zone) and the limnetic zone (open water) in the eastern Chaohu Lake, with three independent sites in each zone (Fig. 1). The eastern Chaohu Lake was divided into grids according to a 10-km2 grid size, and each sampling site was situated in a separate grid. Three limnetic sites were located in the open water zone with water depth of more than 3.5 m. Oppositely, three littoral sites were located in the near lakeshore zone with water depth of less than 2.75 m. Each site was sampled during two seasons, i.e., winter (December 29, 2015) with water temperature ca. 7 °C and autumn (October 19, 2016) with water temperature ca. 22 °C. About 700–1000 g wet weight of mixed surface sediment was collected in duplicate using an Ekman-bottom sampler (Hydro-Bios, Kiel, Germany), placed in sterile polyethylene bags, and stored in an ice cooler before transportation to the laboratory. One part was prefiltered through a 2-mm sieve for chemical analysis; the other was stored at − 80 °C for DNA extraction. Sedimentary temperature and ORP (oxidationreduction potential) were measured during each sampling scenario.
locations of the limnetic zone and pink circles show sampling locations of the littoral zone. The gray curve indicates an area of water depth more than 3 m
J Soils Sediments
2.3 Chemical analysis
2.5 Statistical analysis
To measure the contents of ammonium nitrogen (NH4+-N), nitrate nitrogen (NO3−-N), nitrite nitrogen (NO2−-N), and bioavailable phosphorus (BAP), 5 g fresh sediment sample was extracted with 25 mL KCl (2 mol/L), and placed in an oscillator at room temperature for 2 h (180 r/min). Total phosphorus (TP) was measured by treating at 500 °C (2 h), followed by HCl (1 mol/L) extraction. The above extract was then filtered through a 0.45-μm membrane filter (ANPEL, China) prior to analysis. The concentration of the filtrate was measured using the standard method with a UV-VIS spectrometer (SHIMADZU UV-1700, Japan). Total carbon (TC) and total nitrogen (TN) were performed using an elemental analyzer (ELEMENTAR, Germany) with dried sediment sample passed through a 100-mesh sieve.
Sequences defined as 95% similarity were grouped into one operational taxonomic unit (OTU). OTU-based sequences were aligned using the Clustal X 1.83 program (Thompson et al. 1997). NCBI BLAST (http://www.ncbi.nlm.nih.gov/BLAST) was used to obtain the closest genes sequences in the public database. Neighbor-Joining Trees with bootstrapping were constructed with MEGA 5.0 (Saitou and Nei 1987). Data were expressed as mean ± standard deviation (SD). Two-way analysis of variance (ANOVA) followed by the Student-Newman-Keuls test were used to check for significant differences (p < 0.05) in the physicochemical properties and the gene copy numbers among samples. Correlation analysis was performed using the SPSS v20.0 software. The correlations between each two variables were analyzed using the Pearson correlation coefficient, at the p < 0.05 and p < 0.01 levels of significance. The qPCR data were analyzed using the ABI 7500 System SDS Software (Applied Biosystems, USA). Diversity indices (Shannon and Simpson) and non-parametric richness estimates (Chao1 and Ace) for each clone library were determined using PHYLIP (the PHYLogeny Inference Package) and DOTUR. Graphing was achieved using the Origin 8.5 software.
2.4 DNA extraction, clone libraries construction, and qPCR DNA was extracted from 0.33 g freeze-dried lake sediment sample using FastDNA® Spin Kit for Soil (MP Biomedicals, USA) according to the manufacturer’s protocol. Three replicate DNA extractions of individual samples were combined to produce a single pooled DNA sample and then were checked by electrophoresis on 1% agarose gel. DNA concentration was determined by NanoDrop NANO-2000 UV spectrometer (Thermo Fisher Scientific Inc., USA). The PCR amplification was performed on a T100™ Thermal Cycler (Bio-Rad, USA) using the program as shown in Table S1 (Electronic Supplementary Material). The PCR mixtures contain 1 μL template DNA, 0.25 μL Taq polymerase (TAKARA, Dalian, China), 2.5 μL buffer (10×), 0.25 μL albumin from bovine serum (BAS), 1 μL of each primer paris (synthesized by Sangon Biotechnology, Shanghai, China), 2 μL dNTPs, and finally ddH2O up to 25 μL. The purified PCR products were cloned into a cloning vector and transformed into Escherichia coli. White clones were selected randomly, and the genes of the screened positive clones were sequenced by GENEWIZ, Inc. Quantitative real-time polymerase chain reaction (qPCR) assay was applied to determine the copy numbers of bacterial and archaeal genes. The qPCR amplification was performed on an ABI 7500 FAST (Applied Biosystems, USA) using the SYBR® Premix Ex Taq™ (TaKaRa, Dalian, China) according to the manufacturer’s instructions, with 1 μL DNA extract (1–10 ng) as the template in 25 μL reaction mixtures. Plasmid standards containing the target genes were generated from the appropriate positive clones obtained from clone libraries. The amplification efficiencies were 99% for 16S rRNA and 92–95% for other functional genes. The regression coefficients of standard curves (R2) were both > 0.99.
3 Results 3.1 Physicochemical properties of lake sediments in the littoral and limnetic zones We compared the physicochemical properties of lake sediments in the littoral and limnetic zones at different temperatures (Fig. 2). At the spatial dimension, the sedimentary pH did not show significant difference in both zones, for littoral averagely 6.56 ± 0.19, and limnetic 6.38 ± 0.31. However, we observed significant differences in terms of sedimentary NH4+-N and NO3−-N contents between the littoral zone (61.12 ± 17.52 mg kg−1 and 4.96 ± 1.65 mg kg−1; n = 6) and the limnetic zone (93.82 ± 29.78 mg kg−1 and 6.91 ± 3.22 mg kg−1; n = 6) (p = 0.049 and p = 0.031; n = 6). The spatial variation of TC and OM contents showed a similar trend; in the littoral zone (11.89 ± 1.14 mg kg−1 and 5.32% ± 1.51%; n = 6), levels were lower than in the limnetic zone (13.01 ± 1.49 mg kg−1 and 6.35% ± 0.96%; n = 6) (p = 0.005 and p = 0.013; n = 6). Sedimentary TP values were almost identical in both zones (the littoral zone 638.35 ± 43.64 mg kg−1 and the limnetic zone 638.43 ± 38.78 mg kg−1; n = 6). At the temperature dimension, the average sedimentary ORP in winter (7 °C) (the littoral zone 34.87 ± 3.56 mV and the limnetic zone 48.10 ± 19.73 mV; n = 3) was 2.28 and 1.93 times higher than that in autumn (22 °C) (the littoral zone 17.00 ± 7.21 mV and the limnetic zone 25.00
J Soils Sediments
Fig. 2 Boxplots showed the difference of physicochemical properties of lake sediments in the limnetic zone and the littoral zone at different temperatures. Boxes denote interquartile ranges, round dots denote mean values, and whiskers denote 1.5 times standard deviations. Circle and triangle represent the measured values at different temperatures,
respectively. ORP oxidation-reduction potential, OM organic matters, TC total carbon NH4+-N ammonia nitrogen, NO3−-N nitrate nitrogen, NO2−-N nitrite nitrogen, TN total nitrogen, TP total phosphorus, BAP bioavailable phosphorus
± 7.94 mV; n = 3), respectively. The variation mode of sedimentary ORP in the eutrophic lake followed that of DO in the overlying water column which greatly depends on temperature. In winter (7 °C), the DO concentration in the lake water was 11.27 ± 0.86 mg L−1 (n = 6), with an ORP in the lake sediment of 41.5 ± 14.6 mV (n = 6); in autumn (22 °C), the DO level was 8.38 ± 0.39 mg L−1 (n = 6), with an ORP of 21.0 ± 8.0 mV (n = 6). The NO2−-N contents showed higher values at higher temperatures (0.67 ± 0.23 mg kg−1; n = 6), in general, about two to six times higher than the levels at lower temperatures (0.13 ± 0.02 mg kg−1; n = 6).
Based on the results of the qPCR assay targeting the archaeal amoA gene, AOA abundance ranged from 3.21 × 107 to 4.69 × 108 copies/g. The abundance of AOA showed a tendentious spatial heterogeneity; it was higher in the littoral zone (2.46 × 108 ± 1.55 × 108 copies/g; n = 6) compared with the limnetic zone (1.28 × 108 ± 1.41 × 108 copies/g) (p = 0.07; n = 6). However, it was not driven by temperature (2.36 × 108 ± 1.66 × 108 copies/g at 7 °C vs. 1.39 × 108 ± 1.38 × 108 copies/ g at 22 °C) (p = 0.16; n = 6). The results of correlation analysis showed that AOA abundance was significantly correlated with that of other N cycle-related microorganisms in autumn (22 °C) (Tables 1 and 2). Bacterial amoA gene abundance ranged from 8.11 × 106 to 2.41 × 108 copies/g. The abundance of AOB was significantly lower than that of AOA at each sampling site (p = 0.04; n = 12). Interestingly, AOB abundance showed a high heterogeneity both at both space and temporal (temperature) dimensions (p = 0.04 and p = 0.03, respectively; n = 6), which was not the case for AOA. AOB abundance was strongly correlated with that of AOA (r = 0.950; p < 0.01) and anammox bacteria (r = 0.898; p < 0.05). The abundance of anammox bacteria in the sediments was estimated with the hzsB specific primer sets, and the results showed that there was strong spatial heterogeneity
3.2 Variations of microbial abundance in lake sediments The copy numbers of 16S rDNA genes of total archaea and total bacteria from the lake sediment samples are shown in Fig. 3. The abundance of total archaea and total bacteria ranged from 6.96 × 108 to 6.37 × 109 copies/g and from 2.75 × 109 to 1.94 × 1010 copies/g, respectively. We observed a significant temporal (temperature) heterogeneity, driving the abundance of total archaea (p = 0.056; n = 6) and total bacteria (p = 0.032; n = 6). However, the spatial difference in all samples was not significant (p > 0.05; n = 6).
J Soils Sediments
Fig. 3 Boxplots showed the abundance of N-cycle-related microorganism in lake sediments of the limnetic zone and the littoral zone at different temperatures. Boxes denote interquartile ranges, round dots denote mean values, and whiskers denote 1.5 times standard deviations. Rhombuses and pentagrams represent the measured values at different temperatures, respectively. The percentages above the boxes denote relative abundance of the functional gene. AOA archaeal ammonia monooxygenase subunit
A (amoA) gene, AOB bacterial amoA gene, nirS cytochrome cd1 nitrite reductase gene of denitrifying bacteria, nirK copper-containing nitrite reductase genes of denitrifying bacteria, hzsB hydrazine synthase βsubunit gene of anammox bacteria, nrfA nuclear respiratory factor subunit A gene of DNRA (dissimilatory nitrate reduction to ammonium) bacteria
(p = 0.02; n = 6), which means that anammox bacterial abundance was higher in the littoral zone (9.48 × 106 ± 4.66 × 106 copies/g) than in the limnetic zone (3.74 × 106 ± 4.10 × 10 6 copies/g). Anammox bacterial abundance showed no significant difference between different temperatures (7.95 × 106 ± 5.26 × 106 copies/g at 7 °C vs. 5.27 × 106 ± 5.13 × 106 copies/g at 22 °C; n = 6) (p = 0.19; n = 6). Although both participating in ammonium oxidation, the
abundance values of anammox bacteria were one to two orders of magnitude lower than those of AOA and AOB (Fig. 4). The abundance of the nirS and nirK genes varied greatly with temperature and was higher in winter (7 °C) than in autumn (22 °C). The copy numbers of the nirS gene ranged from 6.09 × 107 to 5.48 × 108 copies/g in winter (7 °C) (n = 6) and from 4.95 × 107 to 2.39 × 108 copies/g in autumn (22 °C)
Table 1 Pearson’s correlations between different kinds of Ncycle-related microbial abundance at 7 °C
amoA of AOA amoA of AOB nirS nirK hzsB nrfA
amoA of AOA
amoA of AOB
nirS
nirK
hzsB
nrfA
1
0.829*
0.847* 0.837* 1 0.839* 0.954** 0.862*
0.602 0.770 0.839* 1 0.803 0.964**
0.931** 0.931** 0.954** 0.803 1 0.758
0.529 0.706 0.862* 0.964** 0.758 1
*
0.829 0.847* 0.602 0.931** 0.529
1 0.837* 0.770 0.931** 0.706
The bold italic entries represent the results described in detail in the text **Correlation is significant at the 0.01 level (two-tailed) *Correlation is significant at the 0.05 level (two-tailed)
J Soils Sediments Table 2 Pearson’s correlations between different kinds of Ncycle-related microbial abundance at 22 °C
amoA of AOA amoA of AOB nirS nirK hzsB nrfA
amoA of AOA
amoA of AOB
nirS
nirK
hzsB
nrfA
1
0.950**
0.976** 0.909* 1 0.974** 0.982** 0.986**
0.912* 0.802 0.974** 1 0.943** 0.938**
0.986** 0.898* 0.982** 0.943** 1 0.985**
0.991** 0.947** 0.986** 0.938** 0.985** 1
**
0.950 0.976** 0.912* 0.986** 0.991**
1 0.909* 0.802 0.898* 0.947**
**Correlation is significant at the 0.01 level (two-tailed) *Correlation is significant at the 0.05 level (two-tailed)
(n = 6). The copy numbers of the nirK gene ranged from 7.22 × 107 to 8.36 × 108 copies/g in winter (7 °C) (n = 6) and from 4.39 × 107 to 1.92 × 108 copies/g in autumn (22 °C) (n = 6). There was no significant difference in the abundance of nirS genes and nirK genes (p > 0.05; n = 12). Correlation analysis indicated a significant positive correlation between nirS and nirK (r = 0.974) (p < 0.01), confirming an interaction with denitrifying bacteria. In terms of N removal processes, the abundances of denitrifying bacteria (nirS 2.55% ± 0.84% of total bacteria; nirK 2.68% ± 1.31% of total bacteria; n = 12) were significantly higher than that of anammox bacteria (0.08% ± 0.03%; n = 12). The abundance of DNRA bacteria, based on nrfA gene detection, showed a tendentious spatial and temporal (temperature) heterogeneity (p = 0.09 and p = 0.07, respectively). Higher transcripts of the nrfA gene were found in the littoral zone (2.41 × 108 to 1.64 × 109 copies/g; n = 6) compared to the limnetic zone (1.94 × 108 to 8.90 × 108 copies/g; n = 6), but were lower in winter (7 °C) (2.34 × 108 to 1.64 × 109 copies/g; n = 6) than in autumn (22 °C) (1.94 × 108 to 6.38 × 108 copies/g; n = 6). Correlation analysis indicated that the abundance of the nrfA gene was significantly correlated with those of the nirS gene (r = 0.986; p < 0.01) and the nirK gene (r = 0.938; p < 0.01). In terms of nitrate-reducing bacteria,
Fig. 4 a Redundancy analysis (RDA) of microbial abundance associated with environmental variables based on qPCR of the six sampling sites at two different temperatures. b Redundancy analysis (RDA) of microbial abundance associated with environmental variables based on qPCR of the
the abundances of DNRA bacteria (7.22% ± 1.91% of total bacteria; n = 12) were significantly higher than those of denitrifying bacteria (p = 0.001; n = 12).
3.3 Microbial community composition in lake sediments Variations in the composition of amoA, nirS, nirK, hzsB, and nrfA genes, involving the four N cycle processes, were observed via clone phylogenetic analysis (Fig. 5). The coverage of clone libraries ranged from 89.11 to 91.06% for samples. For the archaeal amoA gene, a total of 132 sequences were obtained from the 12 samples; a total of 29 operational taxonomic units (OTUs), with a cutoff at 5%, were observed based on the detected archaeal amoA gene in the collected samples. Phylogenetic analysis revealed that the species Nitrosopumilus maritimus, which accounted for 87.1–100.0% of all sequences, was the dominant species in all sampling sites. At the same time, only one OTU containing two sequences, which were recovered from the littoral zone samples, belonged to Nitrososphaera. In terms of AOB, a total of 126 sequences were retrieved from all the 12 samples and could be assigned to 39 OTUs. In contrast to AOA, the AOB community showed a high spatial
six sampling sites in two zones. The samples were marked as Blake region-site number^ or Bseason-site number,^ for example, Blim1^ was one of the sampling site of limnetic zone
J Soils Sediments
Fig. 5 The phylogenetic trees of six functional genes representative sequences (OTUs) from Chaohu Lake sediments. The trees constructed with an alignment of the sequences from OTUs and their closet matched sequences from GenBank. The evolutionary history was inferred using the Neighbor-Joining method. Bootstrap values above 50% of 1000 resamplings are shown near the corresponding nodes. GenBank accession
numbers are shown following each reference sequence. The numbers in the parentheses after each OTU represent the sequence number within this OUT. Six genes were highlight to the six colors, the OTUs in this study were displayed as overstriking words, and the dominant genera were marked as red
heterogeneity (p < 0.01; n = 6); it was dominated by Nitrosospira (80–91.7%) in the littoral zone and by Nitrosomonas (96.5–96.7%) in the limnetic zone. In general, the community diversity of ammonia oxidizers was relatively low in the sediment of Chaohu Lake (Shannon diversity index, 0.82 and 1.58 for AOA and AOB, respectively). In terms of anammox bacteria, a total of 212 sequences were retrieved from the 12 sediment samples and clustered into 64 OTUs. Phylogenetic analysis of the hzsB gene showed that most of the sequences were affiliated with the known anammox bacterial genera, namely Candidatus BBrocadia,^ BScalindua,^ BKuenenia,^ and BJettenia.^ Candidatus
Brocadia (47.1%) and Candidatus Scalindua (49.1%) were the dominant groups in the limnetic zone; Candidatus Jettenia (10.5–11.7%) was always observed at highabundance sampling sites. The denitrifying bacteria in the sediment samples were dominated by Alphaproteobacteria, accounting for 52.5% of the total number of nirS gene sequences and for 52.5% of the total nirK gene sequences. The OTU of the nirS gene with the highest abundance contained 38 sequences and belonged to the genus Bradyrhizobium. Of the nirS gene sequences, 27.6% were affiliated with the class Betaproteobacteria, including Pseudogulbenkiania (7.7%), Azospira (10.5%),
J Soils Sediments
Dechloromonas (5.0%), and Azoarcus (4.4%). Denitrifying bacteria containing the nirK gene were dominated by Mesorhizobium (42.7%), followed by Bosea (21.6%) and Rhizobium (16.4%). For nrfA libraries, the sequences in the limnetic zone were affiliated with Desulfomicrobium (50.0%) and Meiothermus (37.5%). The community in the littoral zone was more abundant and dominated by Desulfomicrobium (55.6%), followed by Anaeromyxobacter (19.0%) and Aeromonas (11.1%).
4 Discussion So far, information on the variation of the abundance and community composition of N cycle-related microorganisms in freshwater sediments between the limnetic and the littoral zone is still limited. The present study is the first integrated report describing the abundance of amoA, nirS, nirK, hzsB, and nrfA genes, involved in ammoxidation, denitrification, anammox, and DNRA processes at different temperatures and in different zones of a large, shallow eutrophic freshwater lake. The abundance values of the six functional genes were lower in the limnetic zone than in the littoral zone, but higher in winter (7 °C) compared to autumn (22 °C). However, for different microorganisms, there were different reasons for such variations. In aquatic systems, AOA and AOB are widely distributed, and in our study, the abundance of AOA was higher than that of AOB. Environmental variation (in terms of oxygen concentration, salinity, pH, and ammonia concentration) most likely affects the niche separation of AOA and AOB (He et al. 2012; Prosser and Nicol 2012; Pett-Ridge et al. 2013). Generally, in a eutrophic lake with high ammonium concentrations, AOA are inhibited by high doses of ammonium, whereas AOB prefer environments with high ammonium concentrations (Verhamme et al. 2011). However, the results from a previous study indicated that under acidic conditions, AOA play a more important role than ammonia-oxidizing bacteria (He et al. 2012). Therefore, the difference between AOA and AOB abundances in this study could be attributed to the low pH (5.83 to 6.82) in the study area. Both AOA and AOB were influenced by spatial variability in nutrients. Chaohu Lake is a eutrophic lake with higher NH4+ contents than the affinity (KS) of AOA (0.0036–0.019 μM NH3) and AOB (4–125 μM NH3) (Martens-Habbena et al. 2009; Martens-Habbena and Stahl 2011). Therefore, NH4+ concentrations may not limit AOA and AOB distribution. Instead, the DO concentration might be the limiting factor. The littoral zone is subjected to varying oxic-anoxic conditions due to water level fluctuations, a continuous exchange between groundwater and lake water, and water level fluctuation (Cirmo and McDonnell 1997; Yin 1999; McClain et al. 2003), providing the required oxygen for the ammonia oxidizers. However, the limnetic zone sediments are usually anoxic. For archaea, the DO
requirement (KS = 3.9 μM O2; Konneke et al. 2005) is lower than that for AOB (KS = 183.3 μM O2; Park and Noguera 2010). This explains the significant spatial heterogeneity in AOB abundance compared to AOA abundance. As mentioned above, the distribution of AOA showed a spatial heterogeneity rather than a temporal (temperature) heterogeneity. Previous studies have found similar results, indicating that there is no significant relationship between AOA abundance and water temperature (Adair and Schwartz 2008). This could be explained by the phylogenetic process of AOA metabolism. Based on the results of the phylogenetic analysis, the AOA community was dominated by N. maritimus, which grows chemolithoautotrophically by aerobically oxidizing NH4+ to NO2− (Konneke et al. 2005). Previously, AOA were widely found in different habitats with temperatures ranging from about − 1 °C (Arctic Ocean; Kalanetra et al. 2009) to 97 °C (hot springs in Iceland; Reigstad et al. 2008). Based on the specific glycerol ether structure of their cell membranes, AOA are adapted to the extreme environmental conditions and keep alive (Liu et al. 2015), which explains why temperature variations from 7 to 22 °C had little effect on the abundance and distribution of AOA. The differences in the spatial distribution of AOB might be attributed to the diverse physiological features of different AOB species. Increased NH4+-N levels favored Nitrosomonas, and the dominant AOB shifted from Nitrosospira in the littoral zone to Nitrosomonas in the limnetic zone. The tolerance level of Nitrosomonas (50–600 mM) to NH4+-N is higher than that of Nitrosospira (39–4500 μM); hence, Nitrosomonas is usually found in ecosystems with high N loads, while Nitrosospira is dominant at low N concentrations (Juretschko et al. 1998; Whitby et al. 2001; Xia et al. 2011). In this study, the NH4+N content in the limnetic zone suppressed Nitrosospira growth, but favored the growth of Nitrosomonas. The distribution of anammox bacteria showed a high spatial heterogeneity, but was not dependent on temperature. After the first report of anammox bacteria in the Black Sea in 2003 (Kuypers et al. 2003), many studies have confirmed that anammox bacteria can survive in a range of extreme environments, with temperatures as high as 85 °C (ridge vent, mid-Atlantic; Byrne et al. 2009) or as low as − 2 °C (sea ice in Greenland; Rysgaard et al. 2004).Otherstudieshaveindicated thatsmall-scaleenvironmental heterogeneity is an important factor shaping the community compositionandabundanceofanammoxbacteria(Zhuetal.2013;Zhu et al. 2015; Zhao et al. 2017). The above discussions could explain that the distribution of anammox bacteria depends on spatial heterogeneity instead of temporal (temperature) changes. Interestingly, Candidatus Jettenia was observed at highabundance sampling sites in this study. This phenomenon was in accordance with the finding in a plant-bed/ditch system of a constructed wetland (Wang et al. 2018). Presumably, the anammox community is capable of quorum sensing (Ding et al. 2013). Candidatus Jettenia can be used to determine the community
J Soils Sediments
compositionandabundanceofanammoxbacteria,mainlybecause it is a facultative anaerobe organism occupying a wide range of niches and, because of its great cell density, it can provide autoinducers (Oshiki et al. 2016). An increase in the relative abundance of some species could be linked to an increased anammox bacteria abundance. Anammox bacteria, AOA, and AOB are all autotrophic organisms involved in the oxidation of ammonium. Aerobic ammonia oxidation provides NO2− as substrate for anammox bacteria, but also competes for NH4+. As a result of the high abundance values (4.91 × 107 to 7.09 × 108 copies/g for AOA + AOB vs. 1.01 × 106 to 1.54 × 107 copies/g for anammox bacteria), ammonium removal processes may be dominated by traditional aerobic ammonia-oxidizing microorganisms in this study area. Denitrifiers depend on nitrifiers, as the latter produce NO3− through microbial oxidation of NH4+ (Prosser 1989). The community composition of denitrifying bacteria differed between the two temperatures and zones in this study. Some studies have indicated that spatial variation in environmental parameters may result in niche partitioning, and this may be a factor in determining the community composition of different nir-type denitrifiers (Philippot et al. 2009; Enwall et al. 2010; Keil et al. 2011). Both denitrification and anammox are N removal processes in aquatic systems. In our study, the abundance of denitrifying bacteria was two–three orders of magnitude higher than that of anammox bacteria. The dominated class of denitrifying bacteria is heterotrophic, requiring organic matter to maintain growth and reproduction while removing nitrate. However, when the ratio of organic matter to NO3− is high, organic matter can inhibit the activity of anammox bacteria (Furukawa et al. 2009). Therefore, N removal processes in the sediment of this eutrophic lake, rich in organic matter, may be dominated by denitrifying bacteria. Another NO3− reduction pathway is DNRA, influencing the cycling and removal of N in aquatic ecosystems. Because of varying environmental parameters, including differences in temperature and nutrient concentrations, the populations of DNRA bacteria differed between sites. A similar study has shown that the dominant DNRA community was influenced by different geochemical and physical parameters in the New River Estuary (Song et al. 2014). Heterotrophic bacteria are relatively sensitive to changes in environmental factors, which was the cause of both spatial and temporal (temperature) heterogeneity. In this study, the abundance of DNRA bacteria was higher than that of denitrifying bacteria. In environments with lower amounts of dissolved oxygen, such as sediments, DNRA bacteria had a greater competitive power (Xie 2016). For the DNRA reaction, 1 mol of NO3− needs to get 8 mol electrons, while only 5 mol electrons are needed for denitrification. Therefore, DNRA bacteria need higher amounts of organic matter as electronic donors than denitrifying bacteria (Tiedje 1988). In our study, DNRA bacteria dominated the NO3− reduction process.
5 Conclusions The present study reveals the significant impacts of temperature and spatial conditions on the distribution of N cyclerelated microorganisms in a large, shallow, eutrophic, non macrophyte-vegetated and turbid lake. The limnetic zone sediment had the higher nitrogen content and the lake littoral zone had the higher microbial abundance and diversity. Furthermore, cloning and sequencing analysis showed that microorganisms were affected by different environmental factors and presented different distribution patterns. The abundance of AOA and anammox bacteria shows spatial heterogeneity. Variations in AOB abundance were also a result of temporal (temperature) heterogeneity. The abundance of denitrifying bacteria and DNRA bacteria varied to a greater extent than temperature. We speculated that the littoral zone was the Bhotspot^ of N-cycling-related microorganisms in a large, eutrophic, and turbid lake. It is suggested that increasing the area and restoring the ecological function of littoral zone was effective and significant in eutrophic lake management. Acknowledgements This research is financially supported by the Major National Water Pollution Control and Management Project (2014ZX07405-003 and 2017ZX07201004-002), National Natural Science Foundation of China (No. 41671471, 41322012 and 51278487), Strategic Priority Research Program of the Chinese Academy of Sciences (XDB15020303), National Key R&D Program (2016YFA0602303), Local Innovative and Research Teams Project of Guangdong Pearl River Talents Program (2017BT01Z176), special fund from the State Key Joint Laboratory of Environment Simulation and Pollution Control (Research Center for Eco-environmental Sciences, Chinese Academy of Sciences) (18Z02ESPCR), Open Research Fund of Key Laboratory of Drinking Water Science and Technology, Chinese Academy of Sciences (16Z03KLDWST), CAS/SAFEA International Partnership Program for Creative Research Teams, and Jiaxing Science and Technology Project (2015AY23008). The author Guibing Zhu gratefully acknowledges the support of a Humboldt Research Fellowship (1152633), and Program of the Youth Innovation Promotion Association (CAS).
Compliance with ethical standards Conflict of interest The authors declare that they have no conflicts of interest.
References Adair KL, Schwartz E (2008) Evidence that ammonia-oxidizing archaea are more abundant than ammonia-oxidizing bacteria in semiarid soils of northern Arizona, USA. Microb Ecol 56:420–426 Ali A, Reddy KR, Debusk WF (1988) Seasonal changes in sediment and water chemistry of a subtropical shallow eutrophic lake. Hydrobiologica 159:159–167 Alvarez-Cobelas M, Velasco JL, Valladolid M, Baltanás A, Rojo C (2005) Daily patterns of mixing and nutrient concentrations during early autumn circulation in a small sheltered lake. Freshw Biol 50: 813–829
J Soils Sediments Bruesewitz DA, Tank JL, Hamilton SK (2012) Incorporating spatial variation of nitrification and denitrification rates into whole-lake nitrogen dynamics. J Geophys Res Biogeo 117:G00N07 Buzzelli CP (1998) Dynamic simulation of littoral zone habitats in lower Chesapeake Bay. I. Ecosystem characterization related to model development. Estuaries 21:659–672 Byrne N, Strous M, Crepeau V, Kartal B, Birrien JL, Schmid M, Lesongeur F, Schouten S, Jaeschke A, Jetten M, Prieur D, Godfroy A (2009) Presence and activity of anaerobic ammoniumoxidizing bacteria at deep-sea hydrothermal vents. ISME J 3:117– 123 Chen YY, Liu QQ (2014) Numerical study of hydrodynamic process in Chaohu Lake. J Hydrodyn 27:720–729 Chen MS, Ding SM, Chen X, Sun Q, Fan XF, Lin J, Ren MY, Yang LY, Zhang CS (2018) Mechanisms driving phosphorus release during algal blooms based on hourly changes in iron and phosphorus concentrations in sediments. Water Res 133:153–164 Cirmo CP, McDonnell JJ (1997) Linking the hydrologic and biogeochemical controls of nitrogen transport in near-stream zones of temperate-forested catchments: a review. J Hydrol 199:88–120 Ding S, Zheng P, Lu HF, Chen JW, Mahmood Q, Abbas G (2013) Ecological characteristics of anaerobic ammonia oxidizing bacteria. Appl Microbiol Biotechnol 97:1841–1849 Ding SM, Chen MS, Gong MD, Fan XF, Qin BQ, Xu H, Gao SS, Jin ZF, Tsang DCW, Zhang CS (2018) Internal phosphorus loading from sediments causes seasonal nitrogen limitation for harmful algal blooms. Sci Total Environ 625:872–884 Enwall K, Throback IN, Stenberg M, Soderstrom M, Hallin S (2010) Soil resources influence spatial patterns of denitrifying communities at scales compatible with land management. Appl Environ Microbiol 76:2243–2250 Francis CA, Roberts KJ, Beman JM, Santoro AE, Oakley BB (2005) Ubiquity and diversity of ammonia-oxidizing archaea in water columns and sediments of the ocean. Proc Natl Acad Sci U S A 102: 14683–14688 Furukawa K, Inatomi Y, Qiao S, Quan L, Yamamoto T, Isaka K, Sumino T (2009) Innovative treatment system for digester liquor using anammox process. Bioresour Technol 100:5437–5443 Haworth EY, Lund JWG, Tutin W (1984) Lake sediments and environmental history. J Ecol 73(2):713 Hayatsu M, Tago K, Saito M (2008) Various players in the nitrogen cycle: diversity and functions of the microorganisms involved in nitrification and denitrification. Soil Sci Plant Nutr 54:33–45 He JZ, Hu HW, Zhang LM (2012) Current insights into the autotrophic thaumarchaeal ammonia oxidation in acidic soils. Soil Biol Biochem 55:146–154 Juretschko S, Timmermann G, Schmid M, Schleifer KH, PommereningRoser A, Koops HP, Wagner M (1998) Combined molecular and conventional analyses of nitrifying bacterium diversity in activated sludge: Nitrosococcus mobilis and Nitrospira-like bacteria as dominant populations. Appl Environ Microbiol 64:3042–3051 Kalanetra KM, Bano N, Hollibaugh JT (2009) Ammonia-oxidizing Archaea in the Arctic Ocean and Antarctic coastal waters. Environ Microbiol 11:2434–2445 Kasuga I, Nakagaki H, Kurisu F, Furumai H (2010) Predominance of ammonia-oxidizing archaea on granular activated carbon used in a full-scale advanced drinking water treatment plant. Water Res 44: 5039–5049 Keil D, Meyer A, Berner D, Poll C, Schützenmeister A, Piepho HP, Vlasenko A, Philippot L, Schloter M, Kandeler E, Marhan S (2011) Influence of land-use intensity on the spatial distribution of N-cycling microorganisms in grass-lands. FEMS Microbiol Ecol 77: 95–106 Konneke M, Bernhard AE, de la Torre JR, Walker CB, Waterbur JB, Stahl DA (2005) Isolation of an autotrophic ammonia-oxidizing marine archaeon. Nature 437:543–546
Kuypers MM, Sliekers AO, Lavik G, Schmid M, Jorgensen BB, Kuenen JG, Damste JSS, Strous M, Jetten MSM (2003) Anaerobic ammonium oxidation by anammox bacteria in the Black Sea. Nature 422: 608–611 Leininger S, Urich T, Schloter M, Schwark L, Qi J (2006) Archaea predominate among ammonia-oxidizing prokaryotes in soils. Nature 442:806–809 Liu S, Hu BL, He ZF, Zhang B, Tian GM, Zheng P, Fang F (2015) Ammonia-oxidizing archaea have better adaptability in oxygenated/hypoxic alternant conditions compared to ammoniaoxidizing bacteria. Appl Microbiol Biotechnol 99:8587–8596 Martens-Habbena W, Stahl DA (2011) Nitrogen metabolism and kinetics of ammonia-oxidizing archaea. Method Enzymol 496:465–487 Martens-Habbena W, Berube PM, Urakawa H, de la Torre JR, Stahl DA (2009) Ammonia oxidation kinetics determine niche separation of nitrifying Archaea and Bacteria. Nature 461:976–979 McClain ME, Boyer EW, Dent CL, Gergel SE, Grimm NB, Groffman PM, Hart SC, Harvey JW, Johnston CA, Mayorga E (2003) Biogeochemical hot spots and hot moments at the interface of terrestrial and aquatic ecosystems. Ecosystems 6:301–312 Megonigal JP, Mines ME, Visscher PT (2005) Linkages to trace gases and aerobic processes. Biogeochemistry 8:350–362 Oshiki M, Satoh H, Okabe S (2016) Ecology and physiology of anaerobic ammonium oxidizing (anammox) bacteria. Environ Microbiol 18: 2784–2796 Park HD, Noguera DR (2010) Characterization of two ammoniaoxidizing bacteria isolated from reactors operated with low dissolved oxygen concentrations. J Appl Microbiol 102:1401–1417 Pett-Ridge J, Petersen DG, Nuccio E, Firestone MK (2013) Influence of oxic/anoxic fluctuations on ammonia oxidizers and nitrification potential in a wet tropical soil. FEMS Microbiol Ecol 85:179–194 Philippot L, Cuhel J, Saby NPA, Cheneby D, Chronakova A, Bru D, Arrouays D, Martin-Laurent F, Simek M (2009) Mapping fieldscale spatial patterns of size and activity of the denitrifier community. Environ Microbiol 11:1518–1526 Prosser JI (1989) Autotrophic nitrification in bacteria. Adv Microb Physiol 30:125–181 Prosser JI, Nicol GW (2012) Archaeal and bacterial ammonia-oxidisers in soil: the quest for niche specialisation and differentiation. Trends Microbiol 20:523–531 Purkhold U, Pommerening-Röser A, Juretschko S, Schmid MC, Koops HP, Wagner M (2000) Phylogeny of all recognized species of ammonia oxidizer based on comparative 16S rRNA and amoA sequence analysis: implications formolecular diversity surveys. Appl Environ Microbiol 66:5368–5382 Reigstad LJ, Richter A, Daims H, Urich T, Schwark L, Schleper C (2008) Nitrification in terrestrial hot springs of Iceland and Kamchatka. FEMS Microbiol Ecol 64:167–174 Robert JN, Henri D (1997) The ecology of interfaces: riparian zones. Annu Rev Ecol Syst 28(1):621–658 Rysgaard S, Glud RN, Risgaard-Petersen N, Dalsgaard T (2004) Denitrification and anammox activity in Arctic marine sediments. Limnol Oceanogr 49:1493–1502 Saitou N, Nei M (1987) The neighbor-joining method: a new method for reconstructing phylogenetic trees. Mol Biol Evol 4:406–425 Schleper C, Nicol GW (2010) Ammonia-oxidising archaea—physiology, ecology and evolution. Adv Microb Physiol 57:1–41 Schubert CJ, Durisch-Kaiser E, Wehrli B, Thamdrup B, Lam P, Kuypers MMM (2006) Anaerobic ammonium oxidation in a tropical freshwater system (Lake Tanganyika). Environ Microbiol 8:1857–1863 Sims A, Horton J, Gajaraj S, McIntosh S, Miles RJ, Mueller R, Reed R, Hu ZQ (2012) Temporal and spatial distributions of ammoniaoxidizing archaea and bacteria and their ratio as an indicator of oligotrophic conditions in natural wetlands. Water Res 46:4121– 4129
J Soils Sediments Song B, Lisa JA, Tobias CR (2014) Linking DNRA community structure and activity in a shallow lagoonal estuarine system. Front Microbiol 5:460 Thompson JD, Gibson TJ, Plewniak F, Jeanmougin F, Higgins DG (1997) The CLUSTAL_X windows interface: flexible strategies for multiple sequence alignment aided by quality analysis tools. Nucleic Acids Res 25:4876–4882 Tiedje JM (1988) Ecology of denitrification and dissimilatory nitrate reduction to ammonium. In: John W (ed) Methods of soil analysis. John Wiley and Sons, New York, pp 179–244 Venter JC, Remington K, Heidelberg JF, Halpern AL, Rusch D (2004) Environmental genome shotgun sequencing of the Sargasso Sea. Science 304:66–74 Verhamme DT, Prosser JI, Nicol GW (2011) Ammonia concentration determines differential growth of ammonia-oxidizing archaea and bacteria in soil microcosms. ISME J 5:1067–1071 Wang SY, Zhu GB, Ye L, Feng XJ, Op den Camp HJM, Yin CQ (2012) Spatial distribution of archaeal and bacterial ammonia oxidizers in the littoral buffer zone of a nitrogen-rich lake. J Environ Sci 24(5): 790–799 Wang SY, Wang WD, Liu L, Zhuang LJ, Zhao SY, Su Y, Li YX, Wang MZ, Wang C, Xu LY, Zhu GB (2018) Microbial nitrogen cycle hotspots in the plant-bed/ditch system of a constructed wetland with N2O mitigation. Environ Sci Technol 52(11):6226–6236 Whitby CB, Hall G, Pickup R, Saunders JR, Ineson P, Parekh NR, McCarthy A (2001) C-13 incorporation into DNA as a means of identifying the active components of ammonia-oxidizer populations. Lett Appl Microbiol 32:398–401 Xia WW, Zhang CX, Zeng XW, Feng YZ, Weng JH, Lin XG, Zhu JG, Xiong ZQ, Xu J, Cai ZC (2011) Autotrophic growth of nitrifying community in an agricultural soil. ISME J 5:1226–1236 Xie BK (2016) Study on the process characteristics and influence factors of disssimilatory nitrate reduction to ammonium by strain desulfovibrio sp. CMX. Dalian University of Technology (in Chinese)
Xing XG, Ding SM, Liu L, Chen MS, Yan WM, Zhao LP, Zhang CS (2018) Direct evidence for the enhanced acquisition of phosphorus in the rhizosphere of aquatic plants: a case study on Vallisneria natans. Sci Total Environ 616-617:386–396 Yang GS, Ma RH, Zhang L, Jiang JH, Yao SC, Zhang M, Zeng HA (2010) Lake status, major problems and protection strategy in China. J Lake Sci 22:799–810 (in Chinese) Yin CQ (1999) The ecological function, protection and utilization of land/ inland water ecotones. J Environ Sci 11:120–124 Yin FC (2011) Study on the evaluation and control strategy of eutrophication in Lake Chaohu. Jia WL, Beijing, pp 35–36 Zan FY, Huo SL, Xi BD, Zhu CW, Liao HQ, Zhang JT, Yeager KM (2012) A 100-year sedimentary record of natural and anthropogenic impacts on a shallow eutrophic lake, Lake Chaohu, China. J Environ Monit 14:804–816 Zhao SY, Zhuang LJ, Wang C, Li YF, Wang SY, Zhu GB (2017) Highthroughput analysis of anammox bacteria in wetland and dryland soils along the altitudinal gradient in Qinghai-Tibet Plateau. Microbiologyopen 7(2):e00556 Zhou LL, Wang SY, Zou YX, Xia C, Zhu GB (2015) Species, abundance and function of ammonia-oxidizing archaea in inland waters across China. Sci Rep 5:159–169 Zhu GB, Wang SY, Wang Y, Wang CX, Risgaard-Petersen N, Jetten MSM, Yin CQ (2011) Anaerobic ammonia oxidation in a fertilized paddy soil. ISME J 5:1905–1912 Zhu GB, Wang SY, Wang WD, Wang Y, Zhou LL, Jiang B, Op den Camp HJM, Risgaard-Petersen N, Schwark L, Peng YZ (2013) Hotspots of anaerobic ammonium oxidation at land-freshwater interfaces. Nat Geosci 6:103–107 Zhu GB, Wang SY, Zhou LL, Wang Y, Zhao SY, Xia C, Wang WD, Zhou R, Wang CX, Jetten MSM (2015) Ubiquitous anaerobic ammonium oxidation in inland waters of China: an overlooked nitrous oxide mitigation process. Sci Rep 5:17306 Zumft WG (1997) Cell biology and molecular basis of denitrification. Microbiol Mol Biol R 61:533–616
