Characterization of Soil Fungal Communities Associated with Native and Invasive Grass Species in Southern Arizona

Project Overview

GW10-030
Project Type: Graduate Student
Funds awarded in 2010: $18,329.00
Projected End Date: 12/31/2010
Region: Western
State: Arizona
Graduate Student:
Major Professor:
Dr. Barry Pryor
University of Arizona

Commodities

  • Agronomic: other, grass (misc. perennial), hay
  • Animals: bovine

Practices

  • Animal Production: grazing - continuous, grazing management, pasture fertility, range improvement, feed/forage
  • Crop Production: conservation tillage
  • Education and Training: on-farm/ranch research
  • Energy: energy conservation/efficiency
  • Farm Business Management: whole farm planning
  • Natural Resources/Environment: biodiversity
  • Pest Management: integrated pest management, weed ecology
  • Production Systems: transitioning to organic, agroecosystems, holistic management, organic agriculture
  • Soil Management: soil analysis, soil microbiology, soil quality/health
  • Sustainable Communities: sustainability measures

    Abstract:

    Loss of native grasslands is fundamentally changing the ecosystem of the arid southwestern United States. These grasslands are threatened by various factors, including the presence of invasive species and environmental perturbations such as declining water tables. In general, plant species are intimately associated with the microbial communities of their environment, with interactions that may range from parasitic to obligate symbiosis. This study investigated the soil fungal communities associated with native grasslands in the southwestern U.S. to determine if community DNA profiling may be indicative of the state of the grassland ecosystem they inhabit. Specifically, the communities associated with Lehmann’s lovegrass (Eragrostis lehmanniana), an invasive grass, and blue grama (Bouteloua gracilis), a sympatric native species, were characterized. In addition, the microbial communities associated with big sacaton (Sporobolus wrightii) growing on sites with varied water tables were also studied.

    Introduction

    One of the most endangered native habitats in the American southwest is the semi-desert grassland. Semi-desert grasslands are characterized by average annual rainfall around 17 inches, with the majority of precipitation falling between April and September (Lowe 1964, Judd 1962). These grasslands are generally found at high elevations, between 5,000 and 7,000 feet (Lowe 1964).

    Semi-desert grasslands have been vastly reduced in range over the past decades, with only 26% of the native grassland of the entire southwestern United States remaining in pristine condition (Cox et al. 1983). Grasslands have been lost entirely due to agriculture, urban development and suppression of fire (Cox et al. 1983). Additionally, many areas that were previously open grassland are now suffering from shrub encroachment which is gradually changing the community structure. An increase in shrub cover increases competition for available resources and alters erosion patterns (Gori & Schussman 2005).

    Las Cienegas National Conservation Area (Las Cienegas NCA) in southern Arizona is a prime example of semi-desert grassland in the United States. The conservation area was established in 2000 and is located approximately 45 miles southeast of Tucson, AZ. The elevation is approximately 4,500 feet above sea level and average annual rainfall is 15 inches, mainly falling during the summer monsoon season. The area is home to a wide variety of plants and animals, including 33 species identified as threatened or endangered (Gori & Schussman 2005). There are over 50 different species of grasses present at the site, including 45 native and eight non-native (invasive) species (Bodner 2009). It is a working cattle ranch and encompasses more than 42,000 acres (Gori & Schussman 2005). Most importantly for this study, over 90% of the area is occupied by native semi-arid grassland (Gori & Schussman 2005).

    Lehmann’s lovegrass (Eragrostis lehmanniana) was originally imported into Arizona from South Africa in 1932 as a forage grass for cattle and to stabilize disturbed soils, and has since been shown to generally outcompete and displace native grasses due to its earlier germination and seeding times, longer growing season and faster growth rate (Cox et al. 1988, Trask 2006). It has bunch type growth and is a warm season C4 grass. Its period of active growth includes spring, summer and fall, with seed production during spring and summer (USDA NCRS PLANT Profile). This grass goes not propagate vegetatively and only propagates from seed. The plant currently can be found throughout much of the southwestern United States, including Arizona, California, New Mexico, Oklahoma, Texas and Utah.

    At Las Cienegas NCA the invasiveness of Lehmann’s lovegrass is particularly destructive because it tends to displace important native grass species such as blue grama (Bouteloua gracilis). Native to the continental United States and Canada, this grass shares its ecological niche with Lehmann’s lovegrass as it is also a warm season C4 grass with bunch growth. In contrast to Lehmann’s lovegrass, blue grama reproduces both by emergence from the seedbank and, more commonly, by vegetative growth from tillers (USDA NCRS PLANT Profile).

    For a species such as blue grama, grazing by cattle also impacts competition because native plants are grazed during the summer growing season whereas Lehmann’s lovegrass, which is not as palatable to livestock, is eaten in fall, winter and spring (Ruyle et al. 1988). In addition, Lehmann’s lovegrass leaves originate at the crown, instead of at the tillers like native grasses, and new leaves elongate horizontally and that protects them from repeated grazing (Cox, Ruyle & Roundy 1990).

    Since its introduction, Lehmann’s lovegrass has invaded an estimated 350,000 acres in the southwest, mostly in southern Arizona (Trask 2006). This grass has been used as forage for cattle, but its long-term presence can destroy native diversity. This invasive species also leads to an increased fire risk, due to a larger amount of standing dried vegetative material during drought conditions and has increased recovery following fire (Brooks & Pyke 2000). Many native plants are not as fire tolerant and quickly disappear once invasives establish. Burning can also have serious negative impacts on soil microbial communities, in particular the fungal communities, which are integral for plant health and soil fertility (O’Dea 2007). Importantly, invasive Lehmann’s lovegrass is capable of robust growth in depleted soils containing little or no beneficial rhizosphere fungi (Eom et al. 1999, Neary et al. 1999).

    Programs for eradication of Lehmann lovegrass are inefficient, requiring substantial funding and years of work. The two main control methods currently in use are chemical control via application of herbicides and hand pulling of the grasses. Herbicides and their application are expensive and kill non-target plants, and hand-pulling is extremely time and labor intensive. Reintroduction of native species has been long considered unfeasible as seeding is unreliable, with very few seeds successfully sprouting (Ethridge et al. 1997). Environmental conditions that precipitate blue grama recruitment events occur from every 30-50 years on silty soils to once every 5,000 years on sandy soils (Lauenroth et al. 1994). The long periods between recruitment events mean that disturbances to the community affecting blue grama grass can have impacts that last for decades. Thus, it is much more economical and reliable to protect the native grasses that are already present from competition from invasive species.

    Protection of the native grasses will improve wildlife habitat, protect local diversity, prevent soil nutrient depletion and lower the risk of wildfires. Control of Lehmann lovegrass will protect other desert species, such as the iconic saguaro cactus, from competition and the deadly consequences of fire. The desert flora and fauna are an integral part of the ecotourism industry of the Southwest, and all are being threatened by these two extremely invasive species. For example, in Tucson, AZ approximately $2 billion in ecotourism dollars per year and almost 40,000 jobs are reliant on preserving the desert landscape (USGS).

    Another important plant community at the Las Cienegas Conservation Area are sacaton flats, a rare plant community type found only in the American southwest. They are defined as containing essentially pure stands of big sacaton (Sporobolus wrightii), interspersed with only a few woody plants. These areas create a unique ecological community that provide habitat and forage for a wide variety of grassland species. Certain species such as the Botteri’s sparrow are specially adapted to live in sacaton and fail to thrive in areas where this plant has been displaced (Jones & Bock 2005).

    Various factors have led to the decline and disappearance of sacaton throughout its native range, including grazing by cattle, development, agriculture and alterations to natural water systems (Gori & Schussman 2005). Big sacaton grows over a wide range of water depths but is most common at water tables between four and six meters deep and is associated with fine textured soils (Stromberg et al. 1996). Water table has been shown to have a strong influence on vegetation composition (Groeneveld and Griepenttrog 1985, Richter 1993, Busch and Smith 1995). Water tables within 3.5 meters of the surface are considered accessible to sacaton plants directly. Plants growing in areas with deeper water tables generally rely on rainfall or floods for moisture (Tiller 2004, Scott et al. 2006). While sacaton plants can survive in such conditions in the long term, they tend not to thrive and new growth is limited (Bodner 2008).

    A previous study investigated the influence of mycorrhizal associations on sacaton seedlings and observed that fungal interactions led to increased growth and survival in both greenhouse and field conditions (Richter & Stutz 2002). Native big sacaton shows significantly improved growth when mycorrhizal fungi are present (Eom et al. 1999, Neary et al. 1999). Conversely, sacaton stand health suffers dramatically in the absence of soil fungi in degraded soils (Eom et al. 1999, Neary et al. 1999, Richter et al. 2002). These data suggest that the health and composition of the members of the soil fungal community may be impacted by the state of the sacaton plant they are associated with. For example, a sacaton plant that is subjected to environmental stresses may not be able to support the same community as an unstressed plant. Conversely, a microbial community that is subjected to environmental stresses may not be able to support a thriving sacaton flat.

    Many grassland ecosystems in Southern Arizona are facing environmental stress in the form of a dropping water tables resulting from a prolonged drought and increased pumping of groundwater for agricultural and urban uses. Along with plant species, most components of the soil biome are negatively impacted by decreasing water availability, and water-stress changes in microbial community structure often occur in advance of subsequent declines in plant health (Miller & Bever 1999).

    It has been observed that changes in water table depth also negatively affect the fungal species present in association with certain grasses such as maidencane (Panicum hemitomon), suggesting that soil fungi may be less tolerant of changes in soil water profiles as their host plants and may serve as an early indicator of degrading conditions (Miller & Bever 1999). It was found that species composition and amount of colonization varied by water depth, even with a constant host plant. The observed differences may have been due to species preference for certain moisture levels for sporulation. For example, they may be able to form associations but not sporulate at high moisture, and some may only sporulate in high moisture conditions (Miller & Bever 1999).

    Stevens and Peterson (1996) also found that the water levels affected the amount and type of mycorrhizal colonization of purple loosestrife. Anderson et al. (1984) also found that fungal species in a mixed plant community were associated with specific environmental conditions, including water level and various soil nutrients. In addition, the same fungal species in different soil conditions formed different associations and varied as to whether these interactions were functional or not.

    Plant-fungal interactions are essential for ecological vitality and stability in virtually all ecosystems, and disruption of this interaction has detrimental effect on both plant health and soil fertility (Gange & Brown 1997). For example, rhizosphere fungi are essential for degradation of plant material, especially cellulose and pectin, which most bacteria are incapable of performing (Newman 1985). The products of this degradation are then recycled for subsequent plant use. Fungi are also responsible for modification of soil structure that facilitates both water retention and plant root penetration (Miller & Jastrow 2000). In a reciprocal manner, plant root exudates then contribute to the soil microbial community by stimulation of specific microbiota, creating a dynamic interaction in symbiosis with the host plant. Conversely, exudates from plant roots may also selectively suppress the growth of deleterious rhizosphere fungi (Sullia 1973). Moreover, soil microorganisms may also interact with the plant directly by performing essential tasks in nutrient uptake that have immediate effects on plant growth and health (Allen 1992).

    With increasing recognition that soil microbial communities are one of the primary components and indicators of ecosystem health and stability, considerable research has recently focused on examining and describing this feature in numerous environments. Traditionally, soil microbial communities have been studied using culturing techniques. Species identification relied heavily on morphology, and non-culturable organisms were not possible to observe and characterize. More recently, the addition of DNA sequencing to the scientific repertoire has allowed for direct DNA sequencing from environmental samples or sequencing of cultured organisms to expand and refine the identification of species from various natural communities. Both culturing and DNA extraction and sequencing are often used in tandem to create a robust profile of microbial communities. However, with these techniques it is difficult to form direct and efficient comparisons between complex samples containing many different species.

    The technique known as Denaturing Gradient Gel Electrophoresis (DGGE) is a well-accepted method for evaluating the diversity of microbial communities (Muzyer & Smalla, 1998). Members of a specific community component can be visualized in a genetic "fingerprint" on a gel, even when the target organisms make up only 1% of the total DNA in the sample (Muzyer et al. 1993). The number of fingerprint bands observed on the gels generally indicates the number of predominant community members (Muzyer & Smalla, 1998). Each band can then be excised, purified and sequenced to identify the exact community member or members it represents.

    Equally important determinant of soil microbial communities are the soil properties themselves. Anderson, Liberta and Dickman (1984) found that plant growth and the presence of mycorrhizal spores was negatively correlated to soil pH, available Ca, Mg, P and soil moisture. Total spore counts were also positively correlated with total N and organic matter. Arbuscular mycorrhizal (AM) fungi can affect the outcome of competition by enhancing resource acquisition of some species of plants over others, and this effect varies based on availability of soil nutrients (Caldwell et al. 1985, Allen and Allen 1990). For example, when phosphorus is abundant, fungal symbiotes can be suppressed (Koide and Li 1990, Koide and Schreiner 1992, DeLucia et al. 1997). AM fungi also affect succession, particularly in cases with both mycorrhizal and non-mycorrhizal species, and they also may increase diversity by favoring less competitive species or vice versa, decreasing diversity by favoring a more competitive species (Allen and Allen 1984, Allen and Allen 1988, Allen et al. 1988, Francis and Read 1994). Therefore, in any comprehensive analysis or comparison of soil microbial communities it is important to evaluate and analyze the substrate on which the community is based.

    Project objectives:

    The objectives of this study were two-fold.

    The first objective was to examine fungal communities associated with a native and an invasive grass species and to develop DGGE fingerprints of the fungal communities associated with each target plant species and then specifically identify key fungi representing each community.

    The second objective was to examine fungal communities associated with big sacaton in Arizona growing on sites with varying water table depths, and to determine if these communities vary by water table, thus providing additional measurable indications as to the stress on these grass communities.

    This data was collected through soil sampling at specific sites and DNA analysis of the resident fungal communities, analysis of physical properties of soil from each site, information on water table subsidence at select sites and correlations between fungal community diversity, grass diversity and environmental features. The data collected may provide an additional metric for grassland health as it relates to pathogenic, commensal and symbiotic relationships at the microbial level and the impact of invasive species on soil fungal communities. Furthermore, by understanding the current state of the microbial communities in soil, we may be able to assess the extent of degradation and predict if and when remedial actions are required or the potential success of such actions.

    The long-term goal of this research is to develop new means to maintain and manage native grass communities and control the spread of invasive grasses. The incorporation of fungal community profiles as an additional means to measure grassland ecosystem health may result in novel measures to make maintenance or restoration programs more efficient and cost effective.

    Any opinions, findings, conclusions, or recommendations expressed in this publication are those of the author(s) and do not necessarily reflect the view of the U.S. Department of Agriculture or SARE.