Spartina alterniflora invasions reduce soil fungal diversity and simplify co-occurrence networks in a salt marsh ecosystem
Graphical abstract
Introduction
Plant invasions are among the most important emerging drivers of ecosystem successions across the world because of globalization and other human-induced environmental changes (Sardans et al., 2017). As a consequence, profound effects on belowground soil microbial communities and the ecosystem processes in which they participate have already been indicated in different ecosystems (Broz et al., 2007; Bradford et al., 2014; Fahey et al., 2020). For instance, soil nitrification rate was altered by the changed soil nitrifying community with the exotic grass invasion (Hawkes et al., 2005). In addition, different fractions of soil microbial communities exhibited distinct responses to plant invasion (Elgersma and Ehrenfeld, 2011). Therefore, to predict the consequences of exotic plant invasions on ecosystems functioning, a better understanding of the interactions between invasive plants and the resident soil microbial community is needed (Graham et al., 2016).
In soil, fungal communities are a quantitatively important microbial group, which maintains the ecosystem balance by contributing to the decomposing soil organic matter and promoting plant growth by facilitating their access to nutrients (Miransari, 2010; Banerjee et al., 2016). They may also act as plant pathogens or, on the contrary, protect the plant against them (Egidi et al., 2019). Due to their close relationship with plants, there is increasing interests in evaluating the response of soil fungal community to exotic plant invasions (Broz et al., 2007; Lekberg et al., 2013) and feedbacks from intrinsic soil fungal communities for plant invasion success (Xiao et al., 2014; Dawson and Schrama, 2016). Invasive grasses (Bromus diandrus and Avena fatua) may lower fungal richness and change the abundance of specific fungal species (such as arbuscular mycorrhizal fungi, AM fungi), which may increase their access to soil nutrients more efficiently than the native plants (Phillips et al., 2019). In contrast, studies with temperate forests revealed an increase in soil fungal community diversity under the invasion with Impatiens glandulifera (Gaggini et al., 2018). Altered fungal communities can even become beneficial to the successful invasions of an exotic plant (Bunn et al., 2015), as shown with Centaurea maculosa acquiring more soil phosphorus in response to changed AM fungi than native grasses (Callaway et al., 2004).
The driving forces of soil fungal diversity and community composition with plant invasion can be a complex result of various active environmental filters and previous land-use history (Zhang et al., 2017a). The importance of environmental filters as drivers may vary with the studied ecosystem types, invasive plant species, and spatiotemporal scales. At the regional scale, environmental filtering was found to being more important than dispersal limitation in structuring soil fungal communities (Kivlin et al., 2014; Zeng et al., 2019). Generally, it is worthy to consider and distinguish the potential activity of soil fungi in order to understand their ecological pattern. All soil fungi are heterotrophic organisms, but they can be divided into three functional groups based on their trophic strategy: (i) pathotroph, (ii) symbiotroph, and (iii) saprotroph (Tedersoo et al., 2014; Nguyen et al., 2016). Pathotrophic fungi may be enriched after plant invasion, thus causing diseases to the native plant, and thereby enhancing the exotic plant invasiveness (Inderjit and van der Putten, 2010). In addition, plant invasion can contribute to releasing fungal pathogens in soils and promoting the establishment of invasive plants (Reinhart et al., 2010). Symbiotrophic fungi (e.g., AM fungi) can provide invasive host-plants with nutrients in exchange for carbon source for their growth (Phillips et al., 2019). Saprotrophic fungi mainly act as organic matter decomposers in soils and regulate carbon and nitrogen flow, which can retain and re-allocate nutrients within their mycelium, and are thereby able to overcome the local nutrient limitation in soils (Boberg et al., 2014; Kyaschenko et al., 2017). Overall, it can be expected that changes in plant-derived carbon sources, as it would occur with an invasive plant can result in the alternation of any of these three soil fungal functional groups.
Fungi may also exhibit quite different preferences for specific edaphic conditions, as determined by soil texture, pH, salinity, and nutrient levels. Previous studies documented that soil nutrient status was the best predictor in determining soil fungal community along a small-scale elevational gradient of alpine tundra (Ni et al., 2018), and a similar result was found in a study across different land-use types (Lauber et al., 2008). Under other environmental conditions and ecosystem properties, soil pH was found as a key underlying driver of soil fungal community composition, e.g. in an Arctic tundra soil (Zhang et al., 2016), in boreal forests after wildfire disturbance (Day et al., 2019), in Andean Yungas forests along an altitudinal gradient (Geml et al., 2014), or even across biomes at a global scale (Tedersoo et al., 2014). Considering conditions as they exist in coastal ecosystems, e.g., mangroves salinity could be an essential environmental filter in shaping rhizosphere fungal community composition (Vanegas et al., 2019).
Coastal ecosystems provide multiple ecological services including mitigation of climate change, biodiversity conservation, sediment and nutrient retention, and water purification (Danovaro and Pusceddu, 2007; Barbier et al., 2011). Plant invasion has been well recognized to pose a serious threat to the sustainability of these ecosystems (Ramus et al., 2017). Smooth cordgrass (Spartina alterniflora Loisel., abbreviated as S. alterniflora), native to the United States, was initially introduced into China in 1979 for coastal protection and eco-engineering purposes (Liu et al., 2016). Since then, it has widely established in coastal mudflats and salt marshes in China along the coastline from Liaoning Province to Hainan Province (Zhang et al., 2017b). In fact, S. alterniflora has become a model plant species for studying invasion ecology in coastal ecosystems. Previous studies have already indicated that S. alterniflora invasion can drastically increase soil organic matter (Cheng et al., 2006; Zhang et al., 2010; He et al., 2019), influence edaphic variables (e.g., pH and salinity) (Zhang et al., 2019), and alter soil nutrients availability (Liao et al., 2007; Xie et al., 2019). Compared with native plant species (Pragmites australis and Scirpus mariqueter), higher nutrient input from S. alterniflora could enhance the resource availability for soil microorganisms (Zhang et al., 2020b), eventually, favor the growth of opportunistic saprotrophic species (Kyaschenko et al., 2017). In recent years, S. alterniflora expanded rapidly in the tidal areas of the Yellow River Delta (YRD), displacing native species and thereby threatening the local aboveground biodiversity (Ren et al., 2019). Yet little is known about the responses of soil fungal diversity and community composition to the S. alterniflora invasion in coastal wetlands.
In this study, we assessed the changes in soil fungal diversity and community after the S. alterniflora invasion in a developing invaded saltmarsh ecosystem. We hypothesize that: (i) soil fungal richness will increase with the S. alterniflora invasion due to increased resource availability triggering a higher abundance of fungal saprotrophs; (ii) the associations among fungal species will become more complex in invaded sites due to the elevated nutrient input with invasion, thus triggering more negative connections as a consequence of species competition for growth.
Section snippets
Study area and soil sampling
This study was conducted in the Yellow-River-Delta National Nature Reserve (37°40′–38°10′N, 118°40′–119°20′E; Fig. 1), located in the Dongying city, Shandong province, China. This area has a warm-temperate continental monsoon climate with four seasons, including a rainy summer. The average annual precipitation is 530–630 mm (Fan et al., 2012) and the annual mean temperature ranges from 11.5 to 12.4 °C (Kong et al., 2015). The exotic S. alterniflora was introduced in the YRD in 1989 to prevent
Edaphic parameters
Soils from the S. alterniflora invaded site had a significantly lower pH and salinity (P < 0.01) (Table 1). In contrast, soil moisture was significantly higher (F1,18 = 15.53, P = 0.001). Soil nutrients content was elevated as a consequence of invasion, as indicated by SOC, DOC, and TN, respectively (P < 0.01). However, there was no statistically significant difference in soil C/N ratios between native and invaded sites (F1,18 = 0, P > 0.05).
Fungal alpha diversity
After quality filtering, a total of 1139 OTUs with a
Effects of plant invasion on soil fungal diversity and community composition
Compared to the native sites, the sites invaded with S. alterniflora showed a remarkably different soil fungal diversity and community composition. High levels of shoot input and litter production from invasive species relative to the native species (Liao et al., 2008; Zhang et al., 2020a) may have promoted the growth of a more diverse fungal community, resulting in a higher number of niches where nutrients would be available (Kubartová et al., 2009; Piper et al., 2015). However, the fungal
Conclusions
Changes in soil fungal diversity, community composition, and functional guilds were investigated in a newly-invaded salt marsh ecosystem in Northern China. Exotic S. alterniflora invasion dramatically decreased fungal richness and phylogenetic diversity compared with native sites with Suaeda salsa. Shifts of the structural and functional fungal community compositions due to S. alterniflora invasion were tightly associated with changes in edaphic parameters, especially soil pH and salinity.
Declaration of competing interest
The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.
Acknowledgements
This study was funded by the National Key Research and Development Program of China (Grant No. 2017YFC0505906), the Fundamental Research Funds for the Central Universities (No. 310430001), and the Interdisciplinary Research Funds of Beijing Normal University. We acknowledge the China Scholarship Council (CSC) for the financial support for Guangliang Zhang to study in Thünen Institute, Germany.
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