FIGURE 8.5. The protective capacity of soil (which governs the silt- and clay-protected carbon and microaggregate-protected carbon pools), the biochemically stabilized carbon pool, and the unprotected carbon pool define a maximum carbon content for soils. The pool size of each fraction is determined by its unique stabilizing mechanism (from Six et al., 2002a).
resources to growth, such as mineral nutrients (Pastor and Post, 1988; Schimel, 1993; Hungate et al., 2003). More holistic modeling efforts of changes in SOM, particularly ones that include soils, plants, herbivores, and detritivores together, are more realistic in their outcomes than those that model the plants, heterotrophs, and soil carbon pools separately (Schimel, 1993; Tinker etal., 1996).
Global change effects on relationships to soil biota (aboveground to belowground) can be modeled as a nested set of control variables.
Morphological features of dominant life forms determine engineering activities at the ecosystem level, physicochemical properties of plant functional groups, modifying the provision of nutritional resources at the community level, and biological properties of individual species controlling direct interactions at the population level (Fig. 8.7) (Wolters et al., 2000). Changes in ecosystem functions created by plant-induced alterations in the disturbance regime, to and by resource consumption rates of soil organisms, should be confined to situations where essential traits of the vegetation are drastically changed. Such a change is most likely when the strength of environmental change overrides all other factors controlling plant assemblage structure, when plants with key attributes or functions invade or become extinct, and when species-poor environments are affected.
Much of the experimental research on ecosystem responses has focused on individual species-level responses, and is seldom concerned with multifactor system-level responses. An ongoing study of ecosystem-scale manipulations in a California annual grassland has now progressed past 3 years' duration. Shaw et al. (2002) measured system-level
responses in control and elevated CO2 plots using the Jasper Ridge Global Change Experiment (JRGCE). The JRGCE imposed four global change factors at two levels: (1) CO2 (ambient and 680 parts per million), (2) temperature (ambient and ambient plus 80 Watts (W) per square meter of thermal radiation), (3) precipitation (ambient and 50% above ambient plus 3-week-long growing season prolongation), and (4) N deposition (ambient and ambient plus 7g of nitrate N per square meter per year), in a complete factorial design. In the third year of manipulations of the JRGCE, elevated CO2 stimulated the production of aboveground biomass in the treatments with all of the other factors at ambient levels. Aboveground biomass increased more than 32%, which is comparable to that in other single-factor CO2 enrichment experiments (e.g., the increase in North Carolina pine plantations was 25%, and the increase in an Arizona free-air CO2 enrichment experiment [FACE] was 20-43% [Kimball et al., 1995]). Interestingly, although each of the treatments involving increased temperature, nitrogen, or precipitation increased aboveground biomass and NPP, elevated CO2 consistently shrank these increases. In fact, with the three other factors and ambient CO2 increased NPP by 84%, but increased CO2 more than halved this, down to 40%. Belowground NPP was even more suppressed, with an average effect across all treatments of minus 22%. This study by Shaw et al. (2002) is an object lesson in the need to pursue multifactorial studies over many years to ascertain the full effects of the manifold variables involved in global change phenomena.
A similar control hierarchy arises in the soil system as follows: attributes of burrowing macroinvertebrates and mammals establish different dynamic equilibria for soil at the ecosystem level; physicochemical properties of trophic groups affect the vegetation at the community level; and biological properties of individual species regulate direct interactions at the population level (Fig. 8.7) (Wolters et al., 2000).
Land use change is probably one of the greatest agents of change in soil biology and ecology, and given the fact that it is so pervasive on all continents now, one can readily agree with Wolters et al. (2000) that land use change rapidly and persistently alters all levels of above- and belowground interactions and acts on a large scale.
Climate change, of the magnitude envisioned over the next century (an average global temperature increase of 2.5°C), will lead to a major shift in the boundaries of ecological systems. There will be climate-induced alteration in the makeup of many plant communities and changes in litter quality due to changes in species composition. For example, as mixed spruce-hardwood forests in the southern boreal region are replaced by hardwood due to global warming, the anticipated higher-quality litter will provide increased availability of resources to the soil organisms. Of course, in more northerly climes, as major climate changes occur, there will be significant alterations in ecosystem functions when it affects organisms that carry out functions performed by few other organisms. Schimel and Gulledge (1998) predicted that in areas where episodic drying and rewetting of soil associated with climate change becomes more severe, populations of cellulolytic and ligninolytic fungi may be reduced, resulting in a decrease in litter decomposition greater than would be predicted by considering only the changes in soil and litter moisture.
ECOLOGY OF INVASIVE SPECIES IN SOIL SYSTEMS: AN INCREASING PROBLEM IN SOIL ECOLOGY
One of the primary concerns of ecology in the 21st century has been the increasing numbers of invasive species in ecosystems around the world. The publicity concerning invasive species in aquatic systems has been extensive. Case studies of lampreys invading the Great Lakes of North America via the Welland Canal and later via the St. Lawrence Seaway, and the rapid spread of the zebra mussel in lakes and streams over much of the Western Hemisphere are noteworthy examples. There is a less obvious but growing literature documenting the effects of introduced plants and animals displacing or outcompeting native species in soils in numerous ecosystems of the world. Some examples follow.
Eastern deciduous forests in North America have been invaded by two species of plants that are often dominant in the understory vegetation. Berberis thunbergii is a woody shrub that often forms dense thickets. Microstegium vimineum, a C4 grass, forms dense carpets. The two invasives co-occur often. In a series of laboratory and greenhouse experiments in New Jersey, Ehrenfeld et al. (2001) found that the soil under these plants was increased in available nitrate and had elevated pH as well. The two invasive plants have different mechanisms to achieve a similar end result. Berberis combines large biomasses of nitrogen-rich roots with nitrogen-rich leaf litter, whereas Microstegium clumps combine small biomasses of nitrogen-rich roots with small biomasses of nitrogen-poor litter that leave much of the surface soil with few roots. Changing key chemical characteristics of soil (e.g., changed nitrate and pH) undoubtedly represent only two of numerous ways in which invasive plant species alter the playing field in contesting for dominance of patches of soil.
In the same research sites that were used by Ehrenfeld et al. (2001), Kourtev et al. (2002) measured alteration of microbial community structure and function by exotic plant species (Japanese barberry [Berberis thunbergii] and Japanese stilt grass [Microstegium vimineum], compared to a co-occurring native species [blueberry—Vaccinium spp.]). They found in both bulk and rhizosphere soils that phospholipid fatty acid (PLFA) profiles, enzyme activities, and substrate-induced respiration (SIR) profiles of microbial communities were significantly altered under the two exotic species. The PLFA profiles provided only an index of community structure rather than specific information about what species were active. A correlation of structure (PLFA) and function, namely enzymes, showed that a particular set of species is associated with a particular pattern of enzyme activities but does not provide information about which of the species were responsible. Kourtev et al. (2002) found that profiles of enzymatic and catabolic capacity in the soil definitely differed with different microbial communities.
One of the more noted plant invasions of the past century was that of the annual grass Bromus tectorum L., which has a current range of 40,000,000 hectares, notably in wide regions of Washington, Oregon, Idaho, and Utah. Evans et al. (2001) measured litter biomass and carbon-nitrogen and lignin-nitrogen ratios to determine the effects on litter dynamics in a site that had been invaded in 1994. Long-term soil incubations (415d) were used to measure potential soil microbial respiration and net nitrogen mineralization. Plant-available nitrogen was measured for 2 years with ion-exchange bags, and potential changes in rates of gaseous nitrogen losses were measured using deni-trification enzyme activity. Bromus invasion significantly increased litter biomass, and its litter had significantly greater carbon-nitrogen and lignin-nitrogen ratios than did native species. The changes in litter quality and chemistry decreased potential rates of nitrogen mineralization in sites with Bromus by decreasing nitrogen available for microbial activity. Evans et al. (2001) suggest that Bromus may cause a short-term decrease in nitrogen loss by decreasing substrate availability and denitrification activity, but over the long term, nitrogen losses are likely to be greater in invaded sites because of increased fire frequency and greater nitrogen volatilization during fire. This mechanism, in conjunction with land use change, will set into play a set of positive feedbacks that will decrease nitrogen availability and alter species composition.
In a companion study to that of Evans et al. (2001), Belnap and Phillips (2001) studied the effects of invasion by Bromus tectorum in three study sites in the Canyonlands of southwestern Utah. They measured litter and soil changes in sites that had been dominated previously by Hilaria jamesii, a fall-active C4 grass, and Stipa comata and Stipa hymenoides, predominantly spring-active C3 species. Belnap and Phillips (2001) measured the abundances of a wide range of microbes, microarthropods, and macroarthropods under Hilaria and Stipa communities, as well as in those that had been invaded by Bromus in 1994 (Fig. 8.8). There were significant changes in numbers and diversity, due in part to changes in amounts and qualities of litter. In the Bromus invaded plots, litter quantity was 2.2 times higher in Bromus and Hilaria together than in Hilaria alone, contrasted with Stipa and Bromus, which was 2.8 times greater than in the Stipa alone. Soil biota responded generally in opposite manners in the plots that combined two perennials and an annual grass. Active bacteria decreased in Hilaria versus Hilaria with Bromus, and increased in Stipa versus Stipa with Bromus. Most higher trophic-level organisms increased in Hilaria plus Bromus relative to Hilaria alone, while decreasing in Stipa plus Bromus relative to Stipa alone. The soil and soil food web characteristics of the newly invaded sites included the following: (1) lower species richness and numbers of fungi and invertebrates; (2) greater numbers of active bacteria; (3) similar species of bacteria and fungi as those invaded more than 50 years previously; (4) higher levels of silt (hence greater water holding capacity and soil fertility); and (5) a more continuous cover of living and dead plant material. The authors note that food web architecture can vary widely from what had existed previously within the same vegetation type, depending on the reactions to the invasive species relative to the previous uninvaded condition. Addition of a common resource can shift conditions significantly, and careful attention to the effects of species by season by site is definitely warranted.
A much different example of a soil invasion is the movement of the predatory New Zealand flatworm, Arthurdendyus triangulatus, into
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