- Fruits: apples, general tree fruits
- Natural Resources/Environment: biodiversity
- Pest Management: biorational pesticides, botanical pesticides, field monitoring/scouting, integrated pest management
- Production Systems: general crop production
- Sustainable Communities: sustainability measures
Biological properties generally responded to all treatment combinations, but tillage provided the strongest treatment effect in most cases. Compared to strip-tillage, moldboard tillage consistently yielded significantly lower values for the following biological measurements: total C and N, above-ground biomass, microbial biomass, enzyme activity, soil respiration, N mineralization, some nematode trophic groups, and earthworms. Compared with organic inputs, synthetic inputs consistently induced significantly lower values for the following biological measurements: microbial biomass, enzyme activity, some nematode trophic groups, and soil respiration. An examination of relationships between biological and physical parameters using redundancy analysis revealed that microporosity was the physical property that was most strongly correlated with most biological parameters. Soil organisms responded to treatments in the following order: tillage > input > rotation.
Tables, figures or graphs mentioned in this report are on file in the Southern SARE office.
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The biological composition of soils is described in terms of biological diversity, the food web, and the richness, abundance, and distribution of species. The soil biological community includes all organisms that live in the soil for at least part of their life cycle and includes species of bacteria, fungi, protozoa, algae, nematodes, arthropods, and worms. Biodiversity of soils is commonly thought of in the context of ecosystems services, or functional roles provided by different subsets of the soil biological community. A selection of the ecosystem services provided by the soil biota includes decomposition of organic matter, and, conversely C retention, nutrient cycling, bioturbation, and suppression of soil-borne diseases and pests (Brussaard et al., 1997). Biological composition is highly variable in soils and dependent upon a number of factors including soil type, moisture content, aeration, organic matter, nutrient availability, temperature, and soil structure.
Ecology is defined as the study of relationships of organisms with each other and their physical environment (adapted from Smiles, 1988). There is no question that organisms are strongly influenced by and also react to their physical habitat. Soil organisms, however, are frequently overlooked despite their critical roles in the functioning of an ecosystem due to their small size, hidden behavior, and uncertainty regarding their contribution to soil processes. Organisms’ responses to changes in their physical and chemical environment are only well enough understood to make broad generalizations.
Soil structure can be evaluated in different ways, but perhaps the most meaningful measurement of structure is an evaluation of the size, configuration, and distribution of soil pores (Danielson and Sutherland, 1986). Soil structure responds to and concurrently exerts a strong influence on the organisms that reside in the soil. The distribution of different pore-size classes and the size of pore openings (i.e. “pore necks”) is often more important for characterizing the soil as a biological habitat than is an evaluation of the soil particles. Precise quantification of soil porosity characteristics such as neck-size and tortuousity is essentially impossible at the present time, due to the complicated nature and opaque fabric of the soil matrix (Danielson and Sutherland, 1986). It is possible, however, to determine pore space with relatively high precision, and by making certain assumptions, the pore-size distribution can be made with at least useful accuracy for both laboratory and field purposes (Danielson and Sutherland, 1986).
Many environmental factors (e.g., climate, soil texture, soil parent material, plant community, management, or season) are critical in determining the composition of soil microbial communities (Bossio et al., 1998). Plants are certainly an important determinant of the soil food web structure. For a given soil, the types and diversity of vegetation are significant determinants of biota diversity and distribution. Plants and their fungal symbionts, cyanobacteria, and algae, serve as the primary producers for all terrestrial ecosystems, and as such control the amounts of C that enter the soil system (Loreau et al., 2001; Wardle, 2002). The variety and complexity of C inputs is influenced by the individual plant species and the assemblage of plant species present in an area (Giller et al., 1997). Since plants respond to their environment with different physiological growth responses, including production of root exudates, growth hormones, and characteristic C:N ratios, the chemical, physical, and biological components of the environment influence the C inputs released to the soil by plants. One may consider the chemical quality of the plant community to be the fiber of which the soil food web is constructed. The primary decomposers of plant litter and the plant pathogen and herbivore species are the foundation upon which the soil food web is spun.
Although plants and other autotrophs are the primary drivers of the soil food web, the heterotrophic organisms which feed directly on the primary producers ultimately dictate the availability of nutrients required for plant growth as well as for the remaining soil microflora and fauna (Sparling, 1985; Jenkinson, 1988). Thus, the plant and decomposer communities are in an obligately mutualistic relationship in which each of the two components relies on the other for their long-term sustenance (Wardle, 2002). The ecology of decomposers is an especially poorly understood area of soil ecology (Neher, 1999) and research results are frequently contradictory among studies.
When considering the primary decomposers of soil organic matter, it is beneficial to remember that microorganisms are responsible for about 90% of the C mineralization resulting from the decomposition of organic compounds (Swift and Anderson, 1993). Other organisms which are also considered primary or secondary decomposers include mites, millipedes, earthworms and termites, and these organisms play an integral role in shredding organic residues and dispersing microbial populations, thereby allowing microorganisms to function more efficiently (Brussaard et al., 1997). The activities of the decomposer species are responsible for most of the biochemical transformations of organic matter, resulting in nutrient mineralization and organic matter humification (Brussaard and Juma, 1996).
Soil microbial biomass is also considered to be a part of the labile soil organic matter, and as such, is an important component of nutrient cycling (Smith, 1979; Paul, 1984). Steenwerth et al. (2002) found in their study of the soil microbial community composition of nine land use types that differences in the community structure of microorganisms were most highly correlated with soil microbial biomass and pH and soil management factors such as fertilizer, herbicide, and irrigation. Bossio et al. (1998) measured phospholipid fatty acid profiles in soils from organic, low-input, and conventional farming systems and ranked the environmental variables governing the composition of microbial communities in their study in the following order: soil type>time>specific farming operation (e.g., cover crop incorporation or side dressing with mineral fertilizer)>management system>spatial variation in the field.
Nematode biodiversity is resistant to reduction of species richness resulting from environmental stressors such as low moisture, chemical biocides, tillage practices, and cropping rotations (Niles and Freckman, 1998; Freckman and Virginia, 1991). Nematode community structure, however, is a sensitive indicator of environmental stressors in terrestrial ecosystems when evaluated at the higher-order (less specific) resolution of nematode feeding groups (Elliott, 1994; Moore and de Ruiter, 1993; Parmalee et al., 1993; Platt et al., 1984; Wardle et al., 1995; Yeates et al., 1991).
Nematodes are a major component of the soil food web in their role as grazers of bacteria and fungi, and thus influence the rate of organic matter decomposition and nutrient turnover. An additional physiological characteristic of nematodes is their metabolic efficiency. Wasilewska et al. (1981) found that bacterial-feeding and fungal-feeding nematodes defecated 80% of the material they consumed in a study of decomposition in a Polish rye field (Secale cerale L.). Depending on environmental conditions, nematodes can directly and indirectly influence organic matter decomposition rates and have been found to strongly enhance the rate of C and nutrient mineralization in some studies (De Ruiter and Moore, 1993; Hunt et al. 1987). Rationale for the use of nematodes in bioassessment efforts is described by Niles and Freckman (1998) as follows: i) environmental chemical stressors are integrated at the organ level in nematodes rather than the cellular level as in bacteria and protozoa; ii) short generation times allow nematodes to respond quickly to many environmental stressors; iii) nematodes are able to survive in polluted soils; iv) because nematodes are limited in their motility to only a few centimeters in soil, they allow the investigation of relatively small spatial areas for environmental growth conditions; v) nematodes influence the productivity and decomposition functions of soil; vi) nematode feeding groups (higher resolutions of biodiversity) inhabit virtually all soils on earth and inflict changes in the soil habitat; and vii) methods for nematode sampling, extraction, and identification are generally straightforward and easy to perform.
Another influential subset of the soil biota is that which alters the physical properties of soils. These organisms are commonly referred to as the soil bioturbators or bioengineers. Earthworms, ants, and termites are the most commonly cited examples of this functional group. Earthworms make the broadest contributions to soil physical property change, especially in agricultural soils. The effects of earthworms on soil structure, formation of heterogenous pores, and increased aggregate stability are hallmarks of soil-biota interactions over long periods of time (Coleman and Crossley, 2003). Both decomposers and bioturbators indirectly influence plant growth (Brussaard, 1997).
Earthworms are widely known to perform many beneficial functions in agricultural systems including shredding residues, enhancing microbial degradation, improving soil fertility, and improving soil physical properties such as water infiltration and aggregate stability. Agricultural management practices strongly impact earthworm populations. The most significantly negative effect on earthworm populations related to agricultural practices is tillage. As the frequency and intensity of tillage operations increase, so does the physical destruction of earthworm burrows, cocoons, and the earthworm bodies themselves (USDA, 2001). Tillage also indirectly affects earthworm abundance by regulating the rate of decomposition and the availability of surface residues. Availability of organic matter is the most important factor related to the soil environment which influences earthworm abundance because organic matter serves as a food source, insulates earthworms from adverse weather, provides shelter from birds and other surface predators, and protects earthworm burrows (USDA, 2001). Studies investigating the effect of different fertilizer sources show that nearly all fertilizers benefit earthworms. Addition of organic sources of nutrients (e.g. manure addition, compost, organic fertilizers) can double or triple earthworm numbers in a single year (Edwards and Bohlen, 1996). The use of inorganic fertilizers has also been found to have a generally positive effect on earthworm numbers, which has been attributed to the indirect effect of increased crop biomass and resulting increases in organic residues (Edwards and Bohlen, 1996; Edwards et al., 1995). In regard to pesticides, most herbicides used today have been found to be harmless to earthworms (USDA, 2001). Insecticides, however, have varying degrees of toxicity for earthworms, depending on the class of chemical used. Insecticides containing carbamates and certain organophosphates are highly toxic to earthworms, as are broad-spectrum fumigants such as fungicides and nematicides (Ernst, 1995; Edwards and Bohlen, 1996).
Agriculture is one of the most obvious examples of how natural soil ecosystems have been drastically altered by humans. The very idea of selectively planting and managing the growth of certain plants for human consumption, and, later, for forage and fiber production, was a hypothesis that was tested by humans in the early ages of our existence and found to be inordinately useful – eventually leading to the agrarian societies our lives today are founded upon. But hundreds of years of artificial manipulation of agricultural soils have dramatically changed the biological and functional capacity of these soils. Giller et al. (1997) expounded on this, pointing out that as agricultural intensification occurs, the biological regulation of soil functions is progressively replaced by regulation through chemical and mechanical means. Therefore, agricultural settings are one of the most logical places to examine the manner in which soil-dwelling organisms respond to calculated changes in their physical and chemical soil habitat.
Soil management factors related to agricultural practices have a wide range of effects on the composition of soil organisms. These practices have both positive and negative consequences with regard to total numbers and diversity of soil organisms. Organisms are affected by agricultural practices both directly and indirectly (Stubbs et al., 2004). Direct effects include bodily damage, habitat destruction, poisoning due to toxic doses of pesticides from target and non-target effects of agricultural chemicals, increased availability of nutrients from fertilizer applications, moisture differences due to evapotranspiration, drainage, and irrigation, and pH differences due to soil amendments. Indirect effects are probably more significant and extensive than direct effects to soil organisms and include soil C reduction due to tillage practices, reduction in complexity and diversity of C inputs due to the reduced diversity of vegetation in the agricultural field, compaction, disturbance of trophic interactions resulting from selective pressure exerted on microbes, and residual toxicity and break-down products of biocide applications (Coleman et al., 1993; Pižl, 1993; Steenwerth et al., 2002; ).
Agricultural management practices such as tillage, residue incorporation, irrigation, and rotation sequence can affect soil microbial biomass (Anderson and Gray, 1990; Sparling et al., 1994; Franzluebbers et al., 1995) as well as soil microbial community composition (Bossio et al., 1998; Lundquist et al., 1999; Calderón et al., 2000). Agricultural practices are recognized to have a strong effect on the functional activity of soil microorganisms (Doran et al., 1987; Coleman et al., 1993). The exact nature of the impacts of agricultural practices such as conventional tillage, monocropping, and chemical fertilizers and pesticides (including fumigation) on the soil microbial community is not well understood and results obtained are often variable in different environments (Capri et al., 1993; Wardle et al., 1999). This reflects the increased complexity of soil ecosystems as compared to air and water systems, for which relatively much more is understood in terms of environmental toxicology (Edwards, 2002). Since soils are one of the two main ultimate sinks for chemical pollutants (the other being water), there is a recognized need to increase our understanding of the effects of conventional agricultural practices on soil microbial communities, as well as how those effects are augmented in subsequent biological trophic interactions, soil fertility, and soil physical properties.
This study investigates three key functional biological groups (microorganisms, nematodes, and earthworms), in a field-scale research experiment originally designed to examine long-term crop yields and pest interactions resulting from three different agricultural management decisions. We selected representative organisms of the microflora, microfauna, and macrofauna classification systems based on relative importance for trophic interactions, nutrient cycling processes, and ability to induce soil physical property change. Body size is commonly used as a functional classification system because it is directly related to metabolic rate, generation time, population density, and food size (Mikola et al., 2002). We construct the composition of the microbial community using taxonomic and functional enumeration techniques. We characterize the nematode and earthworm communities as key representatives of the soil micro- and macrofauna communities that are responsible for significant roles in nutrient cycling and soil physical properties, respectively. Finally, we attempt to establish the linkages between the relative combined presence of these members of the soil community and agricultural practices common to conventional and alternative agricultural systems.
Ecology is defined as the study of relationships of organisms with each other and their physical environment (adapted from Smiles, 1988). There is no question that organisms are strongly influenced by and also react to their physical habitat. As a scientific discipline, soil ecology has only been recognized for two decades and is still frequently overlooked by ecologists (Neher, 1999). This is partially a result of the complexity of soil systems, the broad diversity of soil organisms, the difficulty associated with extracting and identifying soil organisms, and a lack of recognition of the role of soil biota in determining the physical and chemical properties and production potential of soil (Hawksworth and Mound, 1991; Neher, 1999; Loreau et al., 2001). Although soils contain by far the greatest diversity of organisms present in most terrestrial ecosystems, Wardle (2002) estimates that less than 3% of papers published in major ecological journals address studies of belowground organisms. Thus, Wardle and Giller (1996) conclude that the current body of ecological literature and the seminal concepts of terrestrial ecology are based upon only a minority of the Earth’s biota.
The evolution of soil biological communities depends on physical, chemical, and biological soil properties as well as the organisms themselves. The way that these factors come together and affect one another is of great interests to soil ecologists. A conscientious analysis of this topic requires multivariate analytical procedures to quantitatively evaluate relationships and interactions between soil physical properties, management treatments, and soil organisms.
Phospholipid fatty acid (PLFA) analysis is a biochemical method that provides information regarding the soil microbial community composition (Vestal and White, 1989). It may be possible to examine specific PLFAs within a soil to represent indicators of taxonomic or functional groups (Parkes, 1987). Lipids account for 2-20% of the mass of most bacteria, 10-20% of the mass for most fungi, and 2-15% for algae (Jones, 1969; Rattray et al, 1975). Bacteria contain a large diversity of fatty acids including normal, straight-chain, monounsaturated, branched-chain, 2- and 3-hydroxy and cyclopropane configurations (Lechevalier and Lechevalier, 1988). Bacterial polyunsaturated fatty acids are rare (Lechevalier and Lechevalier, 1988). Algae are the microorganisms having the greatest diversity of polyunsaturated fatty acids (Shaw, 1966). Protozoa produce a more limited range of fatty acid compounds and fungi are the most restricted group, most commonly only having 18:2 and 18:3 polyunsaturated fatty acids (Lechevalier and Lechevalier, 1988). Phospholipids, in particular, are apparently always associated with membranes, and certain phospholipids are more common in some groups of microorganisms than in others (Lechevalier and Lechevalier, 1988), which is the principle that makes it useful for characterizing the soil microbial community.
While PLFAs can be useful for examining microbial diversity from a taxonomic perspective, it reveals nothing about the functional diversity of the soil microbial community. Evaluation of enzyme activities corresponding to microbial functions is one way to examine the functional diversity of soils resulting from differing agricultural management systems (Dick et al., 1996; Aon et al., 2001). Fumigation, fertilizer differences, and tillage management are factors in agricultural management that can impact the microorganisms which release many exoenzymes responsible for key soil functions. A reduction in the taxonomic diversity of soils may not impact the functional diversity of soils if there is a high degree of functional redundancy in the soil microbial community (Fonseca and Ganade, 2001; Giller et al., 1997). This occurs because a single soil function is often conducted by a large number of species, so if one of those species is eliminated it has little effect on the function itself because other organisms will fill the functional role (Walker, 1992; Lawton and Brown, 1993). Despite a great deal of time, energy, and money spent researching the consequences of reduced microbial activity resulting from increased agricultural intensity, there is no substantive in situ evidence indicating that microbial diversity is reduced from a taxonomic or functional perspective as a result of intensive agricultural activity (Giller et al., 1997). It is tempting, of course, to predict that soil biodiversity would decrease given that the above-ground diversity of plants and animals is reduced in agriculture fields and, also, that soil disturbance (tillage) often increases as agriculture endeavors become increasingly intensive (Giller et al, 1997). However, chemical pest control substances and tillage both have unpredictable effects on the species diversity of various groups of fauna (Wardle, 1995).
Studies differ in the results achieved by investigating the effect of reduced tillage on pore-size distribution. Some studies have compared pore-size distribution between no-till and conventional tillage and found that macroporosity increased in no-till treatments compared to conventionally plowed treatments (Eliott and Coleman, 1988). Other studies have shown that increases in soil organic matter have variable effects on macro- and microporosity, depending on soil texture. Scheffer and Schachtschabel (1989, translated to English by Kirchmann and Gerzabek, 1999) reported that in fine-textured soils, there was a reduction in macroporosity as a result of reductions in organic matter, but in coarse-textured soils reductions in micropores were observed. Messing et al. (1997) observed that macropores (>75 μm) were formed as a result of higher soil C levels in sandy soils. Azooz et al. (1996) found that the effect of no-till in a silt loam and a sandy loam was to reduce the proportion of large pores and increase the proportion of small pores relative to conventional moldboard plowing. Azooz et al. (1996) concluded that the effect of tillage management on the volume of large pores is specific to the soil and cropping system and further research is required. They suggested evaluating pore size changes due to cropping management under different environmental conditions and soil textures.
This study investigates the way that soil organisms exist and interact with their physical environment. We were fundamentally interested in the soil physical and biological properties resulting from different long-term agricultural management decisions and the relationship between these physical and biological entities. By using phospholipid analysis along with traditional biological activity measurements such as C and N mineralization, we attempted to characterize the soil microbiological community. We extracted the phospholipids present in soil samples from agricultural treatments and compared the phospholipid signatures as indicators of differing microbial community structures resulting from agricultural management decisions. We also evaluated the nematode populations of these agricultural management treatments and related all of these measurements to soil physical properties resulting from the implemented management decisions.
Conservation tillage is broadly defined as any tillage practice that maintains residue cover on at least 30% of the soil surface area (Conservation Tillage Information Center, 1988). Strip-tillage is a conservation tillage practice that isolates soil tillage to a narrow band, generally 15-45 cm in width, using a specialized tractor implement. Strip-tillage incorporates the environmental and crop growth benefits of no-till with the improved root environment associated with tillage practices. Strip-tillage also provides unique soil physical properties compared to conventional or no-tillage practices because the soil matrix undergoes an intermediate level of disturbance relative to these two extreme practices (Hill, 1990; Vyn and Raimbault, 1993).
Positive effects of strip-tillage related to plant growth include factors associated with improved seed bed and rooting environment, decreased surface bulk density, increased moisture content between the rows, soil resistance to negative effects of heavy equipment traffic, and increased yields relative to conventional tillage (Raper et al., 1994). Potential crop benefits related to soil fertility have also been reported and include a more readily mineralizable pool of N and more plant-available P compared to conventional tillage (Kingery et al., 1996). Al-Kaisi and Hanna (2002) report that strip-tillage can improve the seedbed environment in poorly-drained soils due to increases in soil moisture evaporation and increased soil temperature in-row compared to no-till practices.
Other agricultural management practices that have become widely adopted for a variety of reasons are organic production systems and crop rotations. The market for organic produce has seen tremendous growth in the past 15 years. Since 1990, annual growth of organic products has equaled or exceeded 20 percent in retail sales nationally and U.S. certified organic cropland doubled between 1992 and 1997 to 1.3 million acres (Dimitri and Greene, 2002). Interest in rotational cropping has also grown in recent years. In 1997, 82% of the 196 million acres of total U.S. cropland was in some kind of a rotation system (Padgitt et al., 2000). Interest in crop rotations is particularly high for vegetable and alternative crops in the southeast United States as tobacco has become a less economically reliable commodity. Alternative crop management systems offer many advantages to growers and often command a higher price than conventionally-grown crops, especially in the vegetable market. Although it is doubtful that these alternative systems will altogether replace conventional methods of vegetable production in the foreseeable future, there is certainly the potential for these systems to become an accepted and integrated part of the conventional vegetable production system in the southeast United States.
The soil biological community is responsible for many critical crop growth processes including nutrient cycling, soil structure change, and organic matter accumulation/degradation. Therefore, it is important to monitor and assess biological differences between conventional cropping systems and these less-intensive systems (i.e. strip-tillage, organic inputs, and crop rotations). The biological consequences of these alternative management systems have not been studied on a long-term scale in vegetable cropping systems (Hummel et al., 2002).
The extent to which strip-tillage, organic inputs, and crop rotation affect other areas of the field (e.g. the inter-row area between strips) is also unknown. It has been demonstrated that soil biological community composition and biomass varies widely from the rhizosphere to the bulk soil, resulting in significant differences in microbial activity and processes related to soil function, which include soil C sequestration, N dynamics, plant nutrient availability, and litter decomposition (Söderberg and Bååth, 2004). More specific differences in the soil biological activity and composition between a large area of even plant coverage (i.e. inter-rows) and a relatively smaller area of select vegetation (i.e. in-row areas), as seen in strip-tillage systems, is not known. Measurements made for agroecological investigations in conservation tillage studies have been mainly limited to the crop row or crop root zone area. However, since the virtues researchers often extol regarding the advantages of conservation tillage revolve around the fact that reduced tillage systems leave cover crops or previous crop residues on the soil surface, it is of some importance to characterize the ecosystem of the entire field, including the inter-row areas which are frequently cited as the source of environmental and crop-related benefits (e.g. increased C retention, moisture holding capacity, nutrient cycling). It is our impression that previous researchers have treated the inter-row areas of reduced tillage fields as similar to a non-cropped ecosystem, such as a turf grass or pasture system. In the case of strip-tillage, most studies have considered this system as a sort of hybrid between a conventional tillage system (within the strip) and a no-till or turf system (the inter-row). We hypothesize, however, that there may be a radius of influence extending from the perimeter of the strip, causing an ever-decreasing gradient of biophysical effect into the inter-row region.
To evaluate the biological diversity of soils under different agricultural management strategies, recognizing that a highly productive agricultural soil is considered to be one with a high degree of biological activity and containing a stable cross section of microorganisms and invertebrates.
To make the best possible estimations of microbial and invertebrate populations and community structure using a range of direct enumeration and community evaluation techniques.
To assess the effect of microbial and invertebrate communities on the following soil physical properties: aggregate stability, bulk density, porosity, and pore size distribution.
To aid in the assessment of soil degradation by identifying soil biological indicators of high soil productivity potential for agricultural soils.