Biome and Landscape Patterns

Patterns in species richness, food web structure, and functional organization have been observed among biomes and across landscapes. To some extent, patterns may reflect variation in occurrence or dominance of certain taxa in different biomes. Regional species pools may obscure effects of local habitat conditions on species richness (Kozar 1992a), especially in temperate ecosystems (Basset 1996), but few ecologists have addressed the extent to which the regional species pool may influence local species richness. Gering et al. (2003), Kitching et al. (1993), and Progar and Schowalter (2002) distinguished arthropod assemblages among sites within biomes that reflected regional gradients in environmental conditions. Various hypotheses have been proposed to account for apparent patterns at the biome and landscape level (e.g., Price 1997, Tilman and Pacala 1993). However, as more data have become available, some patterns have become equivocal.

General functional groups are common to all biomes (e.g., grazing herbivores [depending on degree of autochthonous primary production in streams], predators, parasites, and detritivores), whereas other functional groups depend on particular resources being present (e.g., sap-suckers require vascular plants and wood borers require wood resources). Proportions of the fauna representing different functional groups vary among biomes. Low-order streams have primarily detrital-based resources, and their communities are dominated by detri-tivores and associated predators and parasites. Other communities represent various proportions of autotroph functional groups (e.g., chemoautotrophs, ruderal, competitive, and stress-tolerant vascular vs. nonvascular plants) and heterotroph functional groups (herbivores, predators, detritivores) (see Chapter 11).

Different species compose these functional groups in different biomes. For example, the insect grazer functional group is composed primarily of moths, beetles, and tree crickets in broadleaved forests, moths and sawflies in coniferous forests (Schowalter 1995, Schowalter and Ganio 1999, Schowalter et al. 1981c), grasshoppers in grasslands and shrublands (Curry 1994), and caddisflies and flies in aquatic communities (e.g., Hart 1992). The predator functional group in terrestrial arthropod communities is dominated by a variety of arachnids, beetles, flies, and wasps, whereas in aquatic arthropod communities this functional group is dominated by dragonflies, true bugs, and beetles.

Among terrestrial biomes, species richness generally is assumed to increase from harsh biomes (e.g., tundra and desert) to grassland to forest, again reflecting differences in physical complexity, suitability, and stability of the habitat (Bazzaz 1975,Tilman and Pacala 1993). However, this trend is not apparent for arthropods among communities where extensive species inventories are available (e.g., Table 9.1). Species richness is not always linearly related to primary productivity and patterns likely depend on scale (Rosenzweig and Abramsky 1993,Tilman and Pacala 1993, Waide et al. 1999). Species richness often declines above intermediate levels of productivity, perhaps because more productive communities are dominated by larger individuals that reduce habitat heterogeneity or because more productive and stable communities favor competitive exclusion of some species by the best adapted species (Tilman and Pacala 1993). For example, continuous fertilization of permanent pasture at Rothamsted, United Kingdom, since 1856 resulted in changes in species rank-abundance pattern from a log normal curve in 1856 to progressively more geometric curves by 1949 (see Fig. 9.3) (Kempton 1979).

Functional group composition has not shown consistent differences among biomes (Hawkins and MacMahon 1989, Stork 1987). Detritivores represent a relatively greater proportion of the community in boreal forests, headwater streams, and other biomes characterized by accumulated organic material and a lower proportion in tropical forests, deserts, and other biomes with little organic matter accumulation (Haggerty et al. 2002, Seastedt 1984). Wood borers occur only in forest or shrub ecosystems with abundant wood resources. Pollinators are more diverse in tropical forests and deserts where plant diversity and isolation have led to greater reliance on insect and vertebrate pollinators, compared to temperate grassland and forest and arctic biomes. Proportional representation of species and individuals among functional groups varies widely among canopy arthropod communities in temperate and tropical forests, depending on tree species composition (Fig. 9.8) (V. Moran and Southwood 1982, Schowalter and Ganio 1998, 1999, Stork 1987).


Western hemlock


Grand fir

□ Parasitoids

□ Detritivores

Western hemlock

Western redcedar

Functional group organization of arthropod communities in canopies of four old-growth conifer species at the Wind River Canopy Crane Research Facility in southwestern Washington. Data from Schowalter and Ganio (1998).

At the landscape or drainage basin scale, patterns in species richness and functional group organization can be related to local variation in physical conditions. The history and geographic pattern of disturbance may be particularly important factors affecting variation in community structure. Polis et al. (1997a) concluded that the movement of organisms and resources among the interconnected community types composing a landscape can contribute to the organization of the broader landscape community by subsidizing more resource-limited local communities. However, Basset (1996) found that diversity in tropical rainforest trees was related to five factors: numbers of young leaves available throughout the year, ant abundance, leaf palatability, leaf water content, and alti-tudinal range. These data suggested that local factors may be more important determinants of local species diversity and community structure in complex ecosystems, such as tropical forests, than in less complex ecosystems, such as temperate forests.

Diversity of stream insects varies among riffle and pool habitats and substrate conditions (Ward 1992). Diversity generally is higher in running water with cobble substrates, with high oxygen supply and heterogenous structure, than in standing water with mud, sand, or gravel substrates.

Vinson and Hawkins (1998) found six studies that compared species richness of stream insects over drainage basins. Species diversity varied with elevation, which co-varied with a number of important factors, such as stream morphology, flow rate and volume, riparian cover, and agricultural or urban land use. In one study Carter et al. (1996) used multivariate analysis (TWINSPAN) to compare species composition among 60 sites representing first-order (characterized by narrow V-shaped channel, steep gradient, nearly complete canopy cover) to sixth-order (characterized by wide channel, low gradient, little canopy cover) streams over a 15,540 km2 drainage basin. They identified five communities distinguished largely by elevation. The highest species richness occurred in mid-order, mid-elevation streams that included species groups characterizing both higher- and lower-order streams.

Transition zones (ecotones) between community types usually have higher species richness because they include species from each of the neighboring communities. Zhong et al. (2003) reported that adult mosquito species diversity was higher at sites surrounded by freshwater and salt marsh than at sites surrounded by either freshwater or salt-marsh alone. Ecotones can move across the landscape as environmental conditions change. For example, the northern edge of Scots pine, Pinus sylvestris, forest in Scotland moved rapidly 70-80 km northward about 4000 years BP then retreated southward again about 400 years later (Gear and Huntley 1991). Sharp edges between community types, such as result from land-use practices, reduce the value of this ecotone as a transition zone.

Patches representing different stages of postdisturbance recovery show distinct patterns of species richness, food web structure, and functional group organization (see Marquis et al. 2002, Chapter 10). Species richness usually increases during community development up to an equilibrium, perhaps declining somewhat prior to reaching equilibrium (e.g., MacArthur and Wilson 1967, E. Wilson 1969). As the number of species increases, the number of species interactions increases. Food chains that characterize simpler communities develop into more complex food webs (E.Wilson 1969). Schowalter (1995), Schowalter and Ganio (1999), and Schowalter et al. (1981c) found that patches of recently disturbed temperate and tropical forests were characterized by higher sap-sucker/folivore ratios than were patches of undisturbed forests, even when data were reported as biomass.

Shure and Phillips (1991) reported that species richness and functional group composition are modified by patch size (see Fig. 6.5). Species richness was lowest in mid-sized canopy openings (0.08-0.4 ha). Herbivore guilds generally had lowest biomass in mid-sized canopy openings; omnivore biomass peaked in the smallest openings (0.016 ha) and then declined as opening size increased; predator biomass was highest in the control forest and smallest openings and lowest in the mid-sized openings; and detritivore biomass was similar among most openings but much lower in the largest openings (10 ha). This pattern may indicate the scale that distinguishes communities characterizing closed-canopy and open-canopy forest. Smaller openings were influenced by surrounding forest, whereas larger openings favored species more tolerant of exposure and altered plant conditions (e.g., early successional species and higher phenolic concentrations) (Dudt and Shure 1994, Shure and Wilson 1993). Intermediate-sized openings may be too exposed for forest species but insufficiently exposed for earlier successional species. However, species richness generally increases with habitat area (Fig. 9.9) (M. Johnson and Simberloff 1974, MacArthur and Wilson 1967) for reasons discussed in the next section.


A number of factors affect community structure (e.g. Price 1997). Factors associated with habitat area, resource availability, and species interactions appear to have the greatest influence.

A. Habitat Area and Complexity

The relationship between number of species and sampling effort, in time or space, has been widely recognized and supported (He and Legendre 2002).The increase in number of species with increasing number of samples reflects the greater representation of the community. Similarly, a larger habitat area will "sample" a larger proportion of a regional species pool (Summerville and Crist 2004). Increasing habitat area also tends to represent increasing heterogeneity of habitat conditions (e.g., M. Johnson and Simberloff 1974, D. Strong et al. 1984), providing an increasing number of niches.

In developing the Theory of Island Biogeography, MacArthur and Wilson (1967) emphasized the relationship between species richness (S) and island area (a), expressed as follows:

where C depends on the taxon and biogeographic region and z is a parameter that varies little among taxa or biogeographic regions, generally falling in the range of 0.20-0.35 (see Fig. 9.9). The value of z increases with habitat heterogeneity and proximity to the mainland. For nonisolated sample areas within islands or within continental areas, the relationship between species number and sample area is similar, but z is smaller, generally 0.12-0.17 (MacArthur and Wilson 1967).

Habitat area has continued to be viewed as a primary factor affecting species richness, likely influencing apparent gradients in species richness with latitude and host residence time (e.g., Birks 1980, Price 1997,Terborgh 1973), as discussed earlier in this chapter. However, habitat area also is a surrogate for habitat heterogeneity. Larger islands are more likely than smaller islands to represent a wider range in elevation, soil types, aspects, etc. Similarly, larger continental areas are more likely than smaller areas to represent a range of habitat conditions. Because relatively distinct component communities develop on particular resources, such as plant or microbial species (e.g., J. Moore and Hunt 1988), species richness increases exponentially as representation of resource diversity increases. Furthermore, habitat heterogeneity provides for refuges from competition or predation (i.e., local patches of competition- or predator-free space). The architectural complexity of individual plants also can affect the diversity of associated fauna (Lawton 1983).

Fragmentation of habitat types often alters species richness and other measures of diversity. Larger fragments retain a greater proportion of species richness than do smaller fragments (Fig. 9.10) (Collinge 2000, Kruess and Tscharntke 2000, Summerville and Crist 2004). Species characteristic of the fragmented habitat often are replaced by species characterizing the surrounding matrix (e.g., Summerville and Crist 2004). Some guilds may be more sensitive to fragmentation than are others. Golden and Crist (1999) reported that sap-sucking herbivores and parasitoids were significantly reduced by fragmentation of a goldenrod community, but chewing herbivores and predators were largely unaffected. Overall insect species richness was reduced by fragmentation, primarily through loss of rare species.

Moth Forst Biome

Forest area (ha)

I3QQ.QJQQ Significant (P < 0.05, R2 = 0.61) relationship between the size of forest fragments and number of woody-plant-feeding moth species in the western Allegheny Plateau of eastern North America. From Summerville and Crist (2004) with permission from Ecography. Please see extended permission list pg 571.

Forest area (ha)

I3QQ.QJQQ Significant (P < 0.05, R2 = 0.61) relationship between the size of forest fragments and number of woody-plant-feeding moth species in the western Allegheny Plateau of eastern North America. From Summerville and Crist (2004) with permission from Ecography. Please see extended permission list pg 571.

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