Vegetation Dynamics and Ecosystem Change
by M.R. George, L.M. Roche and D.J. Eastburn
Vegetation in the annual rangelands (Figure 1) persists as annual dominated grasslands, native dominated grasslands, oak-woodlands, chaparral and coastal scrub. These vegetation types are often adjacent to each other and sometimes mixed in a mosaic. While there is considerable variation in these vegetation types each has botanical characteristics that separate it from the others. In this chapter we will describe the dominant or common species in each vegetation type and we will describe vegetation changes and change agents that are commonly recognized.
The annual rangelands are a mosaic of vegetation types in different states of change and ecosystem services differ for each state. Ecosystem services are the benefits humanity obtains from the environment, and are generally categorized into four service types: provisioning, regulating, habitat, and cultural (TEEB 2010, MEA 2005). For example, chaparral vegetation regulates storm water runoff and reduces erosion adjacent to urban areas, oak-woodlands provide cultural value as open space, and annual grasslands are habitat for several endangered plant and animal species. Because our knowledge of ecosystems services is in its infancy, scientists and managers will need to evaluate ecosystem services as they develop models of vegetation change. At the conclusion of this chapter we will present an approach to evaluating ecosystem services.
In Chapter 2, Ecological History, the authors discussed the evolution of the California grassland, oak-woodlands, chaparral and coastal sage scrub. The purpose of this chapter is to give the reader a broad overview of annual rangeland vegetation and how it may change in response to common disturbances such as invasion, fire and grazing. Other chapters will investigate grazing, livestock and vegetation management practices.
Figure 1. Location and area of annual rangelands (oak woodlands, annual grasslands and chaparral) and other rangeland types in California.
Plant and Animal Communities
Historic Plant Communities
The pre-settlement composition of Mediterranean-type grasslands and the understories of associated shrublands and woodlands, now dominated by non-native annual species, are uncertain. Chapter 2, Ecological History, provides an in-depth review of how the grassland developed and changed over geologic time. Classical ecologist Fredrick Clements first proposed that the vegetation of the Central Valley, the central and southern Coast Ranges, and the valleys of southern California was perennial grassland (Clements 1920) and proposed that these were dominated by Stipa spp. Clements relied on observations of scattered patches of purple needlegrass (Stipa or Nassella pulchra) along railroad rights-of-way (Keeley 1990, Hamilton 1997). It since has been suggested that several other perennial grasses (e.g. Poa secunda, Leymus triticoides, Melica spp., Muhlenbergia rigens) were historically more important constituents in some environments (Keeley 1990, Holland and Keil 1995, Holstein 2001, Schiffman 2007). Recent studies suggest that grasses were dominant only in coastal grasslands and along riparian corridors (Evett and Bartolome 2013) and that grasses were present in the other grasslands but the dominant species may have been native annual forbs (Schiffman 2007, Minnich 2008).
The hypothesis that many of California’s current grasslands were formerly dominated by woody vegetation and not "pristine" prairie (Cooper 1922) has been less popular, but has received some scientific support (Hamilton 1997). Cooper noted numerous examples where repeated burning, often intentionally, was sufficient to eliminate woody vegetation and replace it with weedy annuals. Some annual grassland sites may have previously been dominated by coastal scrub (Hopkinson and Huntsinger 2005) or native annuals (Solomeschch and Barbour 2006 and not perennial bunchgrasses. Keeley (1993) compared site characteristics of grasslands with significant native perennial grass stands and sites lacking native perennial grasses and concluded that in the absence of disturbance by fire and livestock grazing, sites often were re-colonized by shrubs.
While the pre-settlement grassland commonly included native perennial grasses, the composition (species and amounts) of the pre-settlement grassland is uncertain. Invasion of non-native annual species is well documented beginning with European exploration and settlement as early as the late 1600s (Hendry 1931). The major period of invasion was in the 18th century and many of these species were well established by the following century (Keeley 1990) and invasion and expansion continue today.
Current Plant Communities
California’s annual grasslands are generally located below 3000 feet, mostly in a ring around the central valley which is dominated by crop production. Most of California’s grasslands are dominated by non-native grasses and forbs of Mediterranean origin (Heady 1977, Baker 1989, Keeley 1990), although alien taxa in California come from all parts of the world (Hickman 1993).
Plant communities within this ecosystem have not been well defined beyond the classifications of Valley Grasslands and Coastal Prairie. Soft chess brome (Bromus hordeaceus) and broadleaf filaree (Erodium botrys) are common in areas with 65-100 cm (25-40 in) of rainfall, and red brome (B. madratensis) and redstem filaree (E. cicutarium) are common on southern sites with less than 25 cm (10 in) of precipitation (Bartolome et al. 1980). Native perennial grasses are more common on deep soils with high rainfall. Vernal pools, found in small depressions with a hardpan soil layer, support downingia (Downingia spp.), meadowfoam (Limananthes spp.) and other species (Parker and Matyas 1981).
Annual grassland vegetation changes within and between years in response to prevailing weather conditions (see Chapter 1, Mediterranean Climate). Long term change in the annual grasslands and oak-woodland understories are not explained by traditional equilibrium models of plant succession that includes a series of seral stages leading to a stable climax community. Jackson and Bartolome (2002) developed a state and transition model for annual grasslands based on data from nine sites ranging from 16 cm to 160 cm of average annual precipitation in the coast range. They concluded that vegetation change in the annual grasslands is largely controlled by interactions between site and weather. Residual dry matter was important at some sites. They proposed that nonequilibrium models of vegetation change were best for the annual grasslands.
Introduced annual grasses and forbs (Figure 2) dominate the annual grasslands. Soft chess (Bromus hordeaceus, formerly B. mollis), ripgut brome (Bromus diandrus, formerly B. rigidus), wild oats (Avena fatua and A. barbata), red brome (B. madratensis, formerly B. rubens), wild barley (Hordeum spp.), and foxtail fescue (Vulpia myuros) are common grasses (Table 1). Common forbs include broadleaf filaree (Erodium botrys), redstem filaree (E. cicutarium), turkey mullein (Eremocarpus setigerus), true clovers (Trifolium spp.), bur clover (Medicago polymorpha), popcorn flower (Plagiobothrys nothofulvus), and many others. California poppy (Eschscholzia californica), the State flower, is found in the annual grasslands. Native grasses, such as purple needlegrass and blue wildrye (Leymus glaucus) and native forbs can be found throughout the annual grasslands.
Figure 2. Soft chess brome, ripgut brome and wild oats are present in most annual grassland and oak woodland ecosystems in California.
Table 1. Frequency of the 20 most common annual grassland and oak woodland understory species in quadrats along 455 transects located from Mendocino and Shasta Counties to Kern and Ventura Counties (Alonso 2008)
Bromus hordeaceus L.
Bromus diandrus Roth
Avena spp. L.
Lolium multiflorum Lam.
Erodium L'Hér. ex Ait.
Vulpia K.C. Gmel.
Hordeum spp. L.
Bromus rubens L.
Carduus pycnocephalus L.
Trifolium hirtum All.
Taeniatherum caput-medusae (L.) Nevski
Hemizonia spp. DC.
Purple needlegrass NP
Nassella pulchra (A.S. Hitchc.) Barkworth
Purple false brome
Brachypodium distachyon (L.) Beauv.
Bristly dogs tail grass
Cynosurus echinatus L.
Medicago polymorpha L.
Avena fatua L.
Silver hair grass
Aira caryophyllea L.
Spreading hedge parsley
Torilis arvensis (Huds.) Link
Vulpia myuros (L.) K.C. Gmel.
Of the 694 terrestrial vertebrates (amphibians, reptiles, birds, and mammals) native to California, over 285 species utilize annual grasslands for reproduction, feeding and cover, including at least 97 species of mammals, 130 species of birds and approximately 73 species of amphibians and reptiles (CDFG 2011). Some of these species are on state or federal threatened and endangered lists.
Many wildlife species use the annual grasslands for foraging, but some require special habitat features such as cliffs, caves, ponds, or habitats with woody plants for breeding, resting, and escape cover. Characteristic reptiles that breed in annual grassland habitats include the western fence lizard (Sceloporus occidentalis), common garter snake (Thamnophis sirtalis), and western rattlesnake (Crotalus viridis oreganus) (Basey and Sinclear 1980). Mammals typically found in this habitat include the black-tailed jackrabbit (Lepus californicus), California ground squirrel (Spermophilus beecheyi), Botta's pocket gopher (Thomomys bottae mewa), western harvest mouse (Reithrodontomys megalotis), California vole (Microtus californicus), and coyote (Canis latrans) (White et al.1980). The endangered San Joaquin kit fox (Vulpes macrotis mutica) is also found in and adjacent to the annual grasslands (U.S. Fish and Wildlife Service 1983). Common birds known to breed in annual grasslands include the burrowing owl (Athene cunicularia), short-eared owl (Asio flammeus), horned lark (Eremophila alpestris), and western meadowlark (Sturnella neglecta) (Verner et al. 1980). This habitat also provides important foraging habitat for the turkey vulture (Cathartes aura).
Not all Mediterranean grasslands in California are dominated by non-native plants. Vernal pools, serpentine grasslands and coastal prairies, while threatened by alien annual grasses are generally not dominated by them.
Vernal pools are seasonally dry depressions on annual rangelands that support and are typically dominated by native plants. Vernal pools provide habitat for diverse populations of crustaceans, grasses and wildflowers (Solomeshch et al. 2007). Several plants and invertebrates are listed as threatened or endangered under the Federal Endangered Species Act. Vernal pools are threatened by urbanization, agricultural conversion and flood control activities that change the pools hydrology. The invasion of weedy annual species is promoted by hydrologic changes that shorten the period of inundation. Targeted grazing by domestic livestock has proven useful in slowing the invasion of alien annual grasses and maintaining the inundation period (Marty 2005). Trampling by grazing animals may also lengthen the inundation period.
Serpentine grasslands occur on soils that develop on serpentine outcrops mainly in the Franciscan formation of California’s coastal range. Serpentine soils support vegetation that differs from surrounding nonserpentine soils in productivity, floristic composition and often physiognomy. These unique characteristics are the result of low levels of major nutrients (nitrogen, phosphorus and potassium), low levels of calcium combined with high magnesium and high concentrations of potentially toxic elements such as nickel, chromium and cobalt (Harrison and Viers 2007). Many or most plant species (serpentine avoiders) cannot tolerate these harsh conditions but others grow in and out of serpentine (serpentine tolerators). Some species grow only on serpentine. California is one of the world’s hotspots of serpentine endemism. Researchers have identified over 200 species and subspecies that are strongly restricted to serpentine and many others that are weakly restricted.
Serpentine grasslands are best known, not for their grasses, but for their spring wildflower displays including: California goldfields (Lasthenia californica), mixed with many other native forbs in the genera Layia, Gilia, Linancthus, Microseris and Eschscholzia. Purple needlegrass and Sandberg bluegrass (Poa secunda) are often present but not abundant. Alien annual grasses such as wild oats (Avena barbata), rip gut brome (Bromus hordeaceus) and medusahead (Taeniatherum caput-medusae) often become dominant late in the growing season. Invasion of alien annual grasses threatens to displace some serpentine species. Nitrogen enrichment resulting from air pollution has been shown to increase the competitive ability of some annual grasses, thus magnifying their threat to this unique flora. Cattle grazing has been shown to reduce shading from the invasive annuals and is sometimes used to maintain the serpentine flora (Weiss 1999).
Coastal prairie communities grow in discontinuous patches in a thin band near the coast. The prairie is often adjacent to and mixed in a mosaic with coastal scrub communities. The coastal prairie has been separated from the annual dominated valley grassland because of different species composition, and different temperature and precipitation regimes. Keeler-Wolf et al. (2007) described the north coastal grassland that extends from the Oregon border southward to the San Francisco Bay area or further to the central coast depending the ecological investigator.
While the coastal prairie has been invaded by non-native annual grasses and forbs it still is a perennial dominated grassland in most cases. According to Keeler-Wolf et al. (2007) California oat grass (Danthonia californica) is the most important native grass in the coastal prairie. They describe three common types in the oatgrass community. The most common type includes native perennial species such as California brome (Bromus carinatus), purple needlegrass and annual species such as English plantain (Plantago lanceolata), sheep sorrel (Rumex acetosella), subterranean clover (Trifolium subterraneum) and bur clover (Medicago polymorpha). A second type includes tufted hairgrass (Deschampsia caespitosa) and coyote bush (Baccaris pililosa) and a variety of native perennial and exotic annual species. A third type, described as exotic perennial grassland includes hairy oatgrass (Danthonia pilosa) and various rushes (Carex spp) on moist, sometimes acidic soils. Exotic perennial grasses including Kentucky bluegrass (Poa pratensis), velvet grass (Holcus lanatus), sweet vernal grass (Anthoxanthum odoratum), tall fescue (Festuca arundinacea) and Harding grass (Phalaris aquatica) are common in this type.
The coastal prairie is a herbaceous community that has been greatly altered by urbanization and conversion to agriculture. Where these communities remain they have been subjected to invasion of alien annual grasses and forbs similar to that described in the annual grassland section. Livestock grazing, short fire intervals, and drought tend to maintain these grassland communities and limit succession to coastal scrub. Often being adjacent to coastal scrub communities, shrubs such as coyote bush, may invade under long fire intervals and reduced grazing pressure. Ford and Hayes (2007) described coyote bush succession in a state and transition model that included prairie, coyote bush, chaparral and woodland states. Fire intensity and frequency and grazing are the primary disturbances in this state and transition model. Conversion of coastal grassland to coyote bush dominated shrublands has been documented by McBride and Heady (1968) and Russell and McBride (2002)
Historic Plant Communities
The oak woodlands range in structure from a blue oak (Quercus douglasii) savanna with few or no shrubs to a woodland with a shrub layer as well as a tree and understory layer. In a ring around the central valley the blue oak savanna commonly lies between the annual grasslands (lower elevation) and the oak-woodlands (higher elevation).
Most native tree and shrub species are still present in oak-woodland communities but probably in different amounts due to changes in fire frequency, grazing pressure, harvesting and other disturbances. The boundary of the oak savanna and annual grassland may be higher than in the past (see Chapter 2). The species composition of herbaceous vegetation in the oak-woodlands prior to European contact is unknown. It is commonly held that native perennial grasses such as the purple needlegrass were widespread (Clements 1934, Heady 1977). However others have made the case that native forbs were once dominant, especially in drier parts of the woodland (Hamilton 1998). With the introduction of domestic livestock grazing and invasion of alien species during the Spanish colonization, herbaceous cover has changed from perennial to annual and from native to exotic (Holmes 1990). Fire interval and intensity have increased (McClaren and Bartolome 1989). Overstory cover has generally increased (Holzman and Allen-Diaz 1991). Soil moisture late in the growing season has decreased, and soil bulk density has increased due to compaction from large herbivore numbers grazing during the rainy season (Gordon et al. 1989).
Current Plant Communities
The oak woodlands occur in a ring around the central valley and along the coastal mountain ranges. The current oak-woodlands have tree and shrub species composition similar to historic communities but the understory is now dominated by introduced annual grasses and forbs. Native annual and perennial grasses and forbs are present in this annual dominated understory but many are remnants of their former composition. State and transition models for the oak woodlands are based on the fire cycle but may also include vegetation change mediated by disease and grazing (George et al 1992, Eastburn et al. In review).
The oak-woodlands are a multi-layered mosaic of tree, shrub and grassland patches (Figure 3). In some locations these mosaics have been correlated with geological substrate (Cole 1980) and soil characteristics (Harrison et al. 1971). However, other researchers have found each of these vegetation types on most soil depths, slopes, aspects and all geological substrates suggesting that disturbance (fire) and/or biological factors (competition, grazing and browsing) are important determinants of the patchy distribution of these vegetation types (Wells 1962, Callaway and Davis 1991). Given this mosaic of multi-layered vegetation types there is wide amplitude in expected species composition and amounts on the same soil series or association within an ecological site.
Figure 3. The oak-woodlands are often a mosaic of oak, grass and shrub patches.
Oak trees are an important component of the ecosystem serving a valuable role in retention of nutrients which in turn contributes to long-term ecosystem sustainability (Figure 4). Nutrient cycling studies have shown that oak trees create islands of enhanced fertility through organic matter incorporation and nutrient cycling. Compared to adjacent grasslands, soils beneath the oak canopy have a lower bulk density, higher pH, and greater concentrations of organic carbon, nitrogen, total and available P, and exchangeable Ca, Mg, and K (Figure 5), especially in the upper soil horizons (Dahlgren et al. 1997). Removal of oak trees results in loss of soil fertility over a 10 to 20 year period (Kay 1987, Dahlgren et al. 1997).
Figure 4. Nitrogen cycling with major pools of nitrogen (lbs./acre) for an oak woodland-grassland ecosystem in the Schubert watershed at University of California Sierra Foothill Research and Extension Center northeast of Marysville, CA (Dahlgren et al. 2003, California Agriculture 57:42-47).
Figure 5. Selected soil quality and fertility parameters for the 0 to 5 cm surface soils beneath an oak canopy and adjacent grasslands for three oak-woodland sites (Dahlgren et al. 2003, California Agriculture 57:42-47).
While there are around 2000 plant species in the oak-woodlands a few tree, shrub and herbaceous species dominate the species composition. Blue oak, interior live oak (Quercus wislizeni) and coast live oak (Q. agrifolia) are dominants in the oak woodlands (Figure 6). Coast live oak and blue oak are common dominant trees in the coast range. Other trees include toyon (Heteromeles arbutifolia), madrone (Arbutus menziesii) and coffeeberry (Rhamnus californica). The shrub layer, if present, may include narrowleaf goldenbush (Ericameria linearifolia), chamise (Adenostema fasiculatum), black sage (Salvia mellifera), and coast sagebrush (Artemesia californica). In the Sierra Nevada foothills dominant trees include blue oak, interior live oak (Q. wislizenii), and foothill pine (Pinus sabiniana). Black oak (Q. kelloggii) occurs at upper elevations in the transition to coniferous forest. Dominants in the shrub layer, when present, may include wedgeleaf ceanothus (Ceanothus cuneatus), manzanita (Arctostaphylos spp.) and poison oak (Toxicodendron diversilobum). At lower elevations and lower rainfall the oak-woodlands are often an oak savanna. With increasing elevation, rainfall and slope the interior live oak and shrub component increases.
Figure 6. Blue oak, interior live oak and coast live oak are dominant species in the oak woodlands.
The understory is dominated by annual grasses and forbs of European origin. Soft chess brome (Bromus hordeaceus, formerly B. mollis), ripgut brome (Bromus diandrus, formerly B. rigidus) and wild oats (Avena fatua) are the most prevalent grasses in the foothill oak-woodlands and filaree (Erodium spp) is the most prevalent forb. Native perennial grasses such as purple needlegrass and blue wildrye (Leymus triticoides) may also be present. Patches on shallow soils are often dominated by filaree or other low growing forbs. Deep soils with higher water holding capacity are often dominated by wild oats and other tall annual grasses. Oak canopies influence species composition of the understory. Studies have shown that oak canopies favor wild oats, soft chess and ripgut brome (Holland 1980, Ratliff et al. 1991).
Of the 632 terrestrial vertebrates (amphibians, reptiles, birds, and mammals) native to California, over 300 species use oak-woodlands for food, cover or reproduction, including at least 120 species of mammals, 147 species of birds and approximately 60 species of amphibians and reptiles (Tietje et al. 2005). Many of these species are on state and federal threatened and endangered lists.
California quail (Callipepla californicus), Beechey ground squirrels (Spermophilus beecheyi), Botta pocket gopher (Thomomys bottae mewa), are common in the oak woodlands as are Audubon cottontail (Sylvilagus audubonii vallicola), and deer (Odocoileus spp). The rich rodent and lagomorph population is an important food source for common predators including: bobcat (Lynx rufus californicus), coyote (Canis latrans) and the Pacific rattlesnake (Crotalus viridis oreganus). The value of this site for food or cover changes seasonally with the vegetation. In habitat planning each plant community and each species needs must be considered individually and collectively.
Historic Plant Communities
The distribution of chaparral, little changed since the Holocene, is greatest in the Transverse and Peninsular ranges of central and southern California but is also important in the Sierra Nevada foothills and along the coast range. There is little evidence that chaparral is replaced by other vegetation types after a century without fire. Most changes result from changing dominance patterns within the shrub flora. Ceanothus, an obligate fire seeder, varies markedly in its longevity. Some species (e.g. Ceanothus tomentosus) appear to be relatively short-lived, on the order of 30 to 50 years while others persist longer (e.g., C. greggii). Some obligate fire seeders such as manzanita (Arctostaphylos spp) are much longer lived and persist for a century or more (Keeley and Davis 2007).
Current Plant Communities
Chaparral is composed largely of evergreen, sclerophyllous shrub species that range from 1 to 4 meters in height (Figure 7). Other growth forms including soft-leaved subshrubs, perennial herbs, and geophytes (bulbs and corms). Annual herbs are less abundant in mature chaparral but can be present in abundance in early and late successional stands of chaparral (Keeley and Keeley 1984). Sparse stands of trees can occur within chaparral, typically within transition areas with conifers at higher elevations and oaks at lower elevations (Hanes 1977; Keeley and Keeley 1984). Depending on the species composition and underlying topography and soil, the structure of chaparral can range from low, monotonous, smooth-textured vegetation to more heterogeneous stands approaching the vertical structure of woodlands (Keeley 2000).
Figure 7. Chaparral is composed largely of evergreen, sclerophyllous shrub species that range from 1 to 4 meters (3 to 13 feet) in height.
From inland and high elevations to coastal locations, chaparral occurs in both large continuous stands or within a mosaic of vegetation types including coastal sage scrub, annual grasslands, oak woodlands, conifer forests and wetland habitats (Heady 1977; Hanes 1977; Callaway and Davis 1993). Chaparral near the coast tends to occur in disjunct patches occupying more mesic sites whereas coastal sage scrub is distributed more extensively in drier habitats (Kirkpatrick and Hutchinson 1980; Malanson and O’Leary 1994). Mountain foothill and high elevation stands of chaparral are larger and more continuous. Coastal sage scrub occurs in smaller patches generally restricted to steep and south-facing exposures (Keeley 2000; PSBS 1995). Oak woodlands often border chaparral in more mesic areas (e.g., adjacent to stream channels, ravines, north-facing slopes) that have developed deeper soils (Griffin 1977). Oak woodlands are thought to develop within late successional chaparral in areas with more developed soils (Cooper 1922; Wells 1962). The native grassland-chaparral interface is not well understood; however, research has shown cases of type conversion from chaparral to nonnative annual grasslands with frequent fire or mechanical disturbance (Zedler et al. 1983).
The species composition of a particular chaparral stand is largely influenced by fire. Chaparral generally returns to pre-fire structure and composition within a normal fire regime (Keeley 1986); however, considerable research has documented various effects of fire regime on species mortality (Keeley 2000). Frequency of fire has been shown to affect chaparral species composition, where short fire intervals may eliminate obligate seeding species in favor of resprouters (Keeley 1986; 1992). Additional research has shown that fire temperature or intensity also has a strong influence on post-fire species composition (Davis et al. 1989; Rice 1993; Tyler 1995). Stand age following fire is thought to influence the reproduction of species based on reproductive strategies. Research has shown that seedling recruitment is more common for resprouting species in old (> 56 yr.) stands of chaparral whereas seedling recruitment for obligate seeding species was extremely uncommon (Keeley 1986; 1992). This research has led to the conclusion that short interval fires may adversely affect the presence of obligate resprouting species in favor of obligate seeders.
The floristic composition of chaparral varies depending on biogeography, local habitat characteristics and fire history. Of the many growth forms present in chaparral, woody evergreen perennials are the dominant plants and, as such, exert the most influence on the habitat. Chamise (Figure 8) is the most common and widespread species within the chaparral vegetation type (Hanes 1971). This species occurs in most stands of chaparral and is the dominant plant in drier habitats (Keeley 2000). The ubiquity of this species is likely explained by its many adaptations to drought, fire and disturbance (Hanes 1977). Other common shrubs include several species of manzanita and ceanothus, silk-tassel bush (Garrya spp.), oaks (Quercus spp.), redberry (Rhamnus spp.), sumac (Rhus spp.), laurel sumac (Malosma laurina), mountain mahogany (Cercocarpus betuloides), toyon (Heteromeles arbutifolia), holly-leaf cherry (Prunus ilicifolia), and mission manzanita (Xylococcus bicolor) (Holland 1986).
Figure 8. Chamise is the most common and widespread species within the chaparral vegetation type.
Soft-leaved subshrubs are less common in true chaparral than in coastal sage scrub but occur within canopy gaps of mature stands, and may be more prevalent following fire (Holland 1986; Keeley and Keeley 1984; Sawyer and Keeler-Wolf 1995). Common species include California buckwheat (Eriogonum fasciculatum), sages (Salvia spp.), California sagebrush (Artemisia californica), and monkeyflower (Mimulus spp.). Suffrutescent and perennial herbaceous species commonly include deerweed (Lotus scoparius), nightshade (Solanum spp.), Spanish bayonet (Yucca whipplei), rock-rose (Helianthemum scoparium), golden yarrow (Eriophyllum confertiflorum), golden stars (Bloomeria spp.), brodiea (Brodiaea spp.), onion (Allium spp.), and bunch grasses (Nassella spp., and Melica spp.) (Holland 1986; Keeley and Keeley 1984; Sawyer and Keeler-Wolf 1995).
The abundance and diversity of wildlife in California's chaparral is not commonly recognized. The iconic, but now extinct, California grizzly bear (Ursus arctos californicus) and the majestic California condor (Gymnogyps californianus), which nearly became extinct and remains endangered, are the chaparral's most famous animal residents. Chaparral habitat supports nearly 50 species of mammals, but none live exclusively in chaparral. Some are found primarily in mature chaparral and others in young chaparral and along ecotones between chaparral and other plant communities. Several prefer riparian areas in and near chaparral. Predators in California’s chaparral include mountain lions (Puma concolor), bobcats (Lynx rufus) and coyotes (Canis latrans). These predators prey on black tail deer (Odocoileus hemionus columbianus), rabbits and ground squirrels (Quinn 1990).
Although many bird species travel over and through the chaparral, only a few reside year-round. Common birds in chaparral ecosystems include the Wrentit (Chamaea fasciata), Western Scrub Jay (Aphelocoma californica), California Towhee, (Melozone crissalis), Spotted Towhee (Pipilo maculatus) and California Thrasher (Toxostoma redivivum). Birds especially common in chaparral for several years after a fire include Costa's Hummingbird (Calypte costae), Sage Sparrow (Artemisiospiza belli), Rufous-crowned Sparrow (Aimophila ruficeps), Lazuli Bunting (Passerina amoena), Lawrence's Goldfinch (Carduelis lawrencei), and Black-chinned sparrow (Spizella atrogularis) (Quinn 1990).
Postfire succession of birds (Alten 1981, Wirtz 1979a), reptiles (Simovich 1979), mammals (Quinn 1990, Wirtz 1977), and insects (Force 1982) has been studied. Current information suggests that, in general, wildlife habitat may be optimized by maintaining chaparral in many age classes, by restricting the size of burned or treated areas, by protecting trees, and by enhancing water sources (Quinn 1990).
Southern coastal scrub is distributed along the southern and central coast to the South San Francisco Bay Area (Rundel 2007). It extends south into Baja California. The community is sometimes called "soft chaparral" due to the predominance of soft, drought-deciduous leaves in contrast to the hard, waxy-cuticled leaves on sclerophyllous plants of California's chaparral communities. The northern coastal scrub is distributed along the coast from Santa Barbara County to the Oregon border (Ford and Hayes 2007).
Southern coastal scrub on some sites is replaced by chaparral types (Mooney 1977, Gray 1983) but the usual trend of vegetation change in undisturbed stands is toward shrubs of various ages and size classes. Southern coastal scrub is fire-adapted and most species sprout readily from crowns after burning. The coyote brush (Baccharis pilularis) stands in northern coastal scrub have been considered a seral stage in the progression from grassland to woodland or forest (Ford and Hayes 2007).
The extent of southern coastal scrub has been drastically reduced and fragmented by agricultural conversion, urbanization, grazing, altered fire intervals and air pollution (Taylor 2005). While most of the native shrubs remain part of the composition, native annual and perennial grasses and forbs, historically present, have been displaced by the invasion of alien annual grasses and forbs.
Dominant plants of the Northern coastal scrub include Coyote brush (Baccharis pilularis), Yerba Santa (Eriodictyon californicum), coast silk-tassel (Garrya elliptica), salal (Gaultheria shallon), and yellow bush lupine (Lupinus arboreus) are common evergreen shrubs in the northern coastal scrub. Herbaceous species include Western Blue-eyed Grass (Sisyrinchium bellum), Douglas Iris (Iris douglasiana), and several native grasses.
Typical species in the Southern coastal scrub includeCalifornia sagebrush, black sage (Salvia mellifera), white sage (Salvia apiana), California buckwheat (Eriogonum fasciculatum), coast brittle-bush (Encelia californica), and golden yarrow (Eriophyllum confertifolium). Larger shrubs include: Toyon (Heteromeles arbutifolia) and Lemonade berry (Rhus integrifolia). Several native and introduced grasses and forbs are part of this community and cacti and succulents may occur in some locations.
The coastal sage scrub community hosts a great diversity of organisms. Of the many animals that live in the coastal sage scrub, 120 are considered rare, threatened or endangered. Of these, the blue-gray gnatcatcher (Polioptila caerulea) and Stephen’s kangaroo rat (Dipodomys stephensi)are federally endangered. Protection of this unique habitat is critical to the survival of a diversity of animals, including nearly 150 different species of birds and more than 150 different butterfly species (CDFG 2011)
Vegetation Dynamics and Disturbance
The woody and herbaceous plant communities of the annual rangelands are adapted to fire and drought and fire is a major driver of vegetation change. Beginning with European colonization of California invasion of nonnative annual grasses and forbs have had a strong influence on grassland and understory dynamics and on woody plant regeneration. Cultivation, grazing and drought facilitated the invasion of nonnative plants. More recently nutrient enrichment resulting from air pollution has influence herbaceous species composition. Climate change holds the prospect of more change as the amount and timing of rainfall are forecast to change over the coming decades. Conversion of woodlands to grasslands has resulted in permanent, often irreversible changes to woody plant communities. Grazing and browsing have short- and long-term effects on herbaceous and woody plant communities.
Fire strongly influences the structure of annual rangeland plant communities. Lightning caused fires, while infrequent; have surely influenced the structure of these communities. Native Americans used fire as a management tool to enhance habitat and to manage food and fiber plants. While fire is a natural part of annual rangeland ecosystems, fire frequency has changed from frequent burning by Native Americans and early ranchers to infrequent burning today. McClaren (1986) and McClaren and Bartolome (1989) estimated oak woodland fire return intervals of about 25 years prior to European settlement. After settlement the return interval was around 7 years due to burning by settlers. In the 1940s Sampson (1940) estimated that oak-woodland burning by ranchers resulted in return intervals of 8-15 years. While prescribed burning continues today, urbanization and air quality concerns have reduced the use of fire as a management tool. Today fire frequency is more likely to be on the order of 25 to 50 years or longer. Prescribed burning, mechanical and chemical brush control have been used to remove the shrub and tree layers but have been used infrequently since the beginning of the 21st century (Murphy and Crampton 1964, Murphy and Berry 1973).
Historic fire regimes for chaparral are not well-documented, but it appears that the fire return interval was in the range of 50 to 150 years (Conard and Weise 1998). However the fire interval has changed due to anthropogenic ignitions and fire suppression and may be closer to every 50 to 70 years. Minnich (1989) estimated a 70 year fire return interval for chaparral sites in San Diego County but fire interval varies spatially. The fire return interval for coastal shrublands tends to be longer than inland (Keeley and Fotheringham2000). Increasing interval between fires increases the risk of catastrophic fire with far reaching ecological and economic impacts (Allen-Diaz et al. 2007).
Most woody plants in the annual rangelands are either adapted to occasional fire or are able to persist in fire prone ecological regimes. Some resprout following fire from below ground burls (Figure 9), some produce large amounts of dormant seed that persists for long periods of time and are stimulated to germinate by heat or chemical processes initiated by fire and some woody plants exhibit both adaptations (Keely 1977). Live oak trees and chamise resprout following fire while shrubs such as ceanothus are stimulated to germinate by fire. Blue oak tends to be a weak resprouter which contributes to poor regeneration.
Figure 9. Chamise resprouts from the base of the shrub following fire.
Following fire oak woodlands often have a savanna structure until shrubs and small trees begin to fill the space between the existing trees. Competition between the species that germinate or resprout following fire or other disturbances, mediated by weather and soil moisture conditions, greatly influence the vegetation states present in the oak-woodlands. On some soils, geological substrates, and aspects; tree, shrub and grass patches are all possible vegetation states (Figure 10). Shallow soils, coarse and rocky soils and southern aspects sometimes limit vegetation to shrub dominated states. Frequent fire tends to result in vegetation states dominated by an oak-annual grass community (Figure 6.10). Protection from fire and grazing results in a gradual increase in shrubs contributing to increased fuel loads. As the shrub canopy reaches into the tree canopy the potential for crown fires increases (George et al. 1992). Protection from browsing reduces hedging allowing the oak canopy to reach the ground layer increasing the chances for ground fires to become crown fires. Crown fires can top-kill oak trees. While interior live oak (Q. wislizenii) will resprout vigorously, blue oak may not resprout vigorously in some locations. Grazing and browsing may slow the recovery of woody plants following fire (Johnson and Fitzhugh 1990). Vegetation dynamics for many oak woodland sites have been compiled in state and transition models and published by USDA NRCS in ecological site descriptions (http://esis.sc.egov.usda.gov/Welcome/pgESDWelcome.aspx).
Figure 10. Three vegetation states representing early, middle and late in the fire cycle. The early state (PC2.1) is a savannah state with little or no shrub layer immediately following fire. The middle state (PC2.2) shows an increase in the shrub layer but low risk of crown fire and the late state (PC2.3) represents a shrub layer that has grown into contact with the tree layer increasing crown fire risk. Frequent fire tends to result in oak-woodland vegetation states dominated by an oak-annual grass community. Protection from fire and grazing results in a gradual increase in shrubs contributing to increased fuel loads and increased risk of crown fires.
Chaparral undergoes a rapid succession from largely herbaceous flora immediately after fire to relatively dense woody vegetation in a short time period with minimal loss of species (Hanes 1971; Zedler and Zammit 1989). Immediately after a disturbance, usually fire, the grasses and forbs initially dominate. Within 2 - 5 years the seedlings of chaparral plants and the shrubs resprouting from their crown or germinating in response to fire become dominant. Their more aggressive root systems exploit deeper water reserves and they will eventually shade out the forbs and grasses and replace them. By the fifth year shrubs are tall enough to shade out the shorter herbs and approach a climax community.
Early research suggested that without fire, chaparral would develop into oak woodlands or grasslands (Sampson 1944; Wells 1962). Chaparral succession to oak woodlands may occur in mesic situations adjacent to current stands of oak woodlands (e.g., Callaway and D’Antonio 1991) but often examples of greater than 100 year-old chaparral stands show little evidence of further succession (Zedler 1981; Keeley 1992). This research has shown that in addition to remaining stable and reproductively viable following long periods without fire, some chaparral species (most resprouting species) sexually reproduce largely within older aged stands (Zedler 1981; Keeley 1992). Additional research has shown that high frequency burning of chaparral in the presence of non-native grasses can cause type-conversion from shrublands to non-native grasslands (Wells 1962; Zedler et al. 1983; Keeley 1990). So while chaparral appears to be fire-adapted, it can remain healthy for long periods without fire and too-frequent fire may cause conversion to grassland.
Large areas of southern California coastal scrub and chaparral are being altered by the invasion of non-native grasses. Most coastal scrub and chaparral species are adapted to intense but infrequent fire. Under these conditions there is an ephemeral post fire community consisting of annual and perennial herbaceous species that dominate for only one to three years before the shrub canopy closes. With an increase in fire frequency, recruitment of fire-adapted native woody species may be hindered, slowing the formation of a closed woody canopy. Under these conditions non-native grasses and other herbaceous species persist longer after fire, and grasses may dominate patches in mature coastal sage scrub and chaparral communities. However, the exact role of non-native grasses during recovery of these plant communities from fire remains unclear (Pec and Carlton 2014). Deposition of oxidized nitrogen associated with urban pollution sources appears to strengthen the competitive ability of non-native grasses. Therefore, the relationship between invasive grasses, fire, and loss of coastal scrub appears to be exacerbated by nitrogen deposition that increases exotic grass biomass more rapidly than native plants (Weiss, 1999, Allen et al., 2005, Fenn et al., 2010 and Kimball et al., 2014).
Volatilization of nitrogen and to a lesser degree potassium are important fire associated nutrient losses. Some nitrogen is recovered or replaced by nitrogen fixing legumes such as lupine (Lupinus spp.) and deerweed (Lotus spp.), and non-leguminous plants such as California lilac (Ceanothus spp.). The interrelationships among soil microorganisms, heating rates associated with wildfires or prescribed burns, soil moisture at the time of a fire, and various nitrogen-fixing plant species have been studied, but much remains to be learned about the dynamics of nutrients in chaparral systems. Soil erosion following fire results in large losses of all nutrients (Conrad et al. 1986).
Successional changes to southern coastal scrub following fire are complex and may vary with geographic region as well as fire interval, intensity and seasonality. Some shrub species are strong resprouters that regrow and flower during the first growing season following a fire while other species are weak- or non-sprouters that are more dependent on germination and seedling establishment and recover more slowly. In the coyote bush dominated northern coastal scrub long intervals between fire and removal of livestock grazing facilitate succession to a shrub dominated community (McBride and Heady 1968, and Russell and McBride 2002). Livestock grazing and frequent fire can maintain a grassland. Without fire and grazing northern coastal scrub expands into unmanaged fragments of land resulting from agricultural, urban and industrial conversion.
In the annual grassland and woody understory fire reduces thatch build up and grass dominance resulting in a shift in species composition toward forb dominance. When fire occurs its effect is short lived. The first growing season after a fire, forage production will commonly be reduced by about 25 to 50 percent partly because species composition commonly is dominated by filaree and other forbs (Hervy 1949, Stromberg and Kephart 1996). The grass component is usually recovering by the second growing season following a fire and by year three species composition and productivity is back to pre-fire levels.
The annual grasslands and the understories of annual rangeland woody plant communities are commonly dominated by annual grasses and forbs that invaded during the European colonization of California. While yearly and within year variation in productivity and species composition is heavily influenced by prevailing weather, long-term change in annual grassland and oak woodland and shrub understory productivity, species composition and ecosystem processes has been influenced by continuing waves of invasion (DiTomaso et al. 2007). Structural changes in invaded plant communities typically cause reduced native species richness and diversity and changes in canopy structure.
In the annual grasslands invasive plants have altered ecosystem structure and function including hydrologic, fire and nutrient cycles. Replacement of deep rooted native perennial grasses by annual grasses and forbs that are largely rooted in the top 12 inches (30 cm) of the soil has changed patterns of soil moisture depletion leaving a soil moisture niche for invading summer annuals such as yellow starthistle (Holmes and Rice, Dyer and Rice 1999, Gerlach 2004). Use of deeper soil moisture by yellow starthistle may mimic use of deeper soil associated with former native plants. Additionally loss of deep rooted perennials has reduced the transfer of nutrients stored below 12 inches to the surface soil.
Non-native grasses and forbs, being prolific seed producers, have displaced most of the native perennial seed bank resulting in extremely high seedling densities following fall germination. The high rates of self-thinning and turnover that follow potentially result in a large flux of nitrogen as these seedlings decompose (Eviner and Firestone 2007). This may contribute to high N cycling rates in exotic annual grasslands.
Alien annual grasses have been shown to reduce oak seedling growth and survivability by limiting soil moisture. Non-native grasses and forbs compete with seedlings of woody plants by depleting soil moisture at more rapid rates than perennials, especially in early spring when acorns are germinating and sending down their roots. Rapid soil moisture depletion rates in annual dominated understories are devastating to oak seedlings compared to more gradual depletion rates of perennial dominated understories (McCreary 2001).
Grazing animals consume forage, redistribute nutrients and compact soil. However the influence of these processes on vegetation change are not well documented (Jackson and Bartolome 2007) and are strongly influenced by prevailing weather. Within the strong constraints exerted by prevailing weather, grazing can influence short and long-term vegetation change. Grazing managers can manage these effects by controlling the season, intensity, frequency, duration and distribution of grazing. Chapter 8 (Grazing Management) reviews the influence of these principles of grazing and their short term effect on the annual grasslands and oak-woodland understory. Grazing effects on the annual grassland or herbaceous understory, where most plants are annual and short lived, tend to be short term while grazing effects on longer lived woody species tend to be longer in term.
Grazing effects on the herbaceous component of annual rangelands tends to reduce grass dominance and thatch build up that reduces light availability to the forb component of the composition. This same effect occurs with the removal of litter or RDM by mowing. However, prevailing weather strongly influences production and species composition that results from manipulation of the grazing process (George et al. 2001).
Grass dominance and thatch build up that results from removal of grazing can have devastating effects on forbs, some of which are critical habitat for insects or other animals. Weiss (1999) found that dwarf plantain (Plantago erecta), a forb that is critical habitat for the bay checkerspot butterfly (Euphydryas editha bayensis), is suppressed by competition from invasive grasses. Removal of grazing from this habitat can be devastating to the checkerspot caterpillars.
Strategically applied livestock grazing has the potential to engineer vegetation structures that meet the habitat needs of endangered animal species such as the San Joaquin Kit Fox (Vulpes macrotis) and the tiger salamander (Ambystoma tigrinum). For example the San Joaquin Kit Fox prefers a relatively open habitat and often disappear from ungrazed habitats. Barry et al. (2011) have reviewed grazing impacts and strategies that can be used to manipulate habitat for several animal species in California’s annual rangelands.
Trampling, especially during the wet season, may result in soil compaction the effects are not uniform. Studies at the San Joaquin Experimental Range in Madera County have shown that compaction increases with moderate or heavy grazing when compared to no grazing (Tate et al. 2004). Other researchers have also found increased soil bulk density in grazed compared to ungrazed pastures (Liacos 1962, Ratliff and Westfall 1971, and Assaeed 1982). When grazing is removed bulk density may decrease over a period of several years (Tate et al. 2004).
While grazing effects on annual rangeland nutrient dynamics have not been observed, it is generally accepted that grazing animals accelerate nutrient cycling by bypassing the decomposition pathway. However, nutrient redistribution is not uniform because livestock distribution is not uniform. Tate et al. (2000, 2003) found that livestock deposit excreta in patches reflecting their preferential use of a pasture landscape.
Grazing has positive and negative effects on oak-woodland ecosystem sustainability. Positive grazing effects include reduced moisture competition between oaks and herbaceous understory, reduced habitat for rodents that consume oak seedlings and acorns and elimination of ladder fuels that increase the risk of crown fire. Negative effects of grazing include increased soil compaction due to grazing during the wet season, consumption of acorns and oak seedlings and reduced soil organic matter (McCreary 2001, Allen-Diaz et al. 2007).
Burrowing animals including the ground squirrels (Spermophilus beecheyi), gophers (Thomomys bottae) and voles (Microtus californicus) can have a dramatic effect on annual rangeland productivity and species composition. Ground squirrels and gophers disturb huge amounts of soil throughout most of the annual rangelands. Disturbed mounds are excellent microsites for germination and establishment of annual seedlings (Stromberg and Griffin 1996, Dyer and Rice 1997). Seed predation may also be an important effect of annual rangeland rodent populations. Voles and house mice (Mus musculus) have been shown to decrease wild oat numbers compared to foxtail barley and ripgut brome (Borchert and Jain (1978).
Disease may have a role in vegetation change. Barley yellow dwarf virus, transmitted by aphids infects many introduced and native grasses in the annual grasslands and oak-woodland understories and reduces survivorship and seed yield in some species (Malstrom 1998). Crown rust has been found to reduce biomass and reproduction in wild oaks (Carsten et al. 2001). Sudden oak death is a new disease affecting oaks in California and Oregon. It is caused by Phytophthora ramorum which is a newly described pathogen. Plant species that are not killed by this disease act as a reservoir for the pathogen.
Because scientists are uncertain whether climate change will result in warmer, cooler, wetter or drier conditions the effect of climate change on individual plant performance and the structure of plant communities is uncertain. Warmer, wetter weather and elevated atmospheric CO2 may increase productivity and result in changes in species composition. These changes may also cause some plant communities to increase in size and extent and others to be reduced. Observed and predicted rise in CO2 may facilitate invasions by nonnative plants. Future species composition and structure of annual rangeland communities will be determined by a suite of global changes potentially resulting in new dominant species and new community structures (Dukes and Shaw 2007, Shaw et al. 2011).
Because the effects of climate change cannot be determined by experimentation alone, scientists use global models to predict future precipitation and temperature patterns. Climate studies indicate that on average California ecosystems will experience warmer, wetter winters, and slightly warmer summers but there is no evidence that the seasonal Mediterranean climate will change. The winters will remain wet and cool and the summers dry and hot. However the spatial and temporal distribution of winter precipitation, the frequency of extreme events and the length of the growing season may change. These changes in temperature and precipitation will result in changes in ecosystem structure, function and services (Dukes and Shaw 2007).
While the full extent of climate change impacts on rangeland forage production and species composition in California’s annual rangelands is not extensively studied, one study forecasts changes in precipitation patterns on California rangeland production and concluded that areas of the state suitable for cattle grazing would shift, as some areas become wetter and others become drier, depending on the climate model. Statewide, they predicted range forage production would decline between 14 and 58 percent (Shaw et al. 2011).
Scientists predict that if climate change results in warmer temperatures, lower humidity, higher winds and drier fuels, fire ignition rates and spread will increase. Torn et al. (1998) forecasted that climate change will result in increased number of fires that escape containment in regions that have large amounts of grass or brush fuels.
The distribution of vegetation may change in response to global climate change. Because some of California’s oaks are constrained by climatic factors, some scientists have hypothesized that their range may be reduced and their location may shift to the north (Kueppers et al. 2005). Scientists studying life history strategies in California’s Mediterranean shrublands hypothesize that climate change trends toward warmer winter temperatures will favor facultative sprouters and increasing rainfall will favor nonsprouters and obligate resprouters, while reduced precipitation will favor facultative sprouters. Increasing fire probability will favor facultative species, while decreasing fire probability will favor obligate resprouting species. Because future climatic and fire regimes may favor one life history strategy over another, the distribution of shrub species and communities may change (Ramirez et al. 2012).
Ecosystem services are the benefits humanity obtains from the environment, and are generally categorized into four service types: provisioning, regulating, habitat, and cultural (TEEB 2010; MEA 2005). California’s annual grasslands and oak woodlands provide multiple benefits to society, including forage and livestock production, wildlife habitat, recreation, carbon sequestration, and drinking water supply (Table 6.2). Management and conservation of rangelands is critical in maintaining ecosystem function and capacity to support goods and services over time. Services can be provided locally by an ecosystem, but the benefits to human well-being can also accrue across multiple scales (de Groot et al. 2010). For example, agricultural production can provide food at the local and global levels; managed watersheds and open space provide water and nutrient cycling and community value at the regional level; and conservation practices can provide carbon sequestration and climate regulating functions at the global level.
Across California’s annual grasslands and oak woodlands, there has been a historical focus on agricultural production, with the goal of sustaining the state and national food supply; however, there is increasing societal demand for provisioning agricultural goods (e.g., livestock and forage production) and additional services (e.g., abundant and high quality water, wildlife habitat) through the management and conservation of these lands (Briske 2011). Balancing tradeoffs between agricultural production and the maintenance of ecosystem services will be a key challenge. Here, we highlight an example framework for understanding multiple ecosystem service provisioning across a managed oak woodland-annual grassland system.
During the mid-20th century, approximately 1.9 million acres of oak woodland were cleared to create productive, open grasslands (Biswell 1954; Murphy and Crampton 1964; Bolsinger 1988). The UC Sierra Foothill Research and Extension Center (SFREC)—located in the northern Sierra Nevada foothills in Yuba County, California—has been a natural laboratory for oak woodland research (McCreary 2010). At SFREC, woody species (predominantly Q. douglasii, Q. wislizeni, Ceanothus spp., and Toxicodendron diversilobum) were actively cleared during the 1960s for forage improvement objectives, and selective woody species removal continued throughout the 1970s and late 1980s. The resulting gradient of woody cover (i.e., cleared open grassland, thinned savanna, and unthinned woodland) has served as a model managed landscape to assess tradeoffs and synergies between multiple ecosystem service-based goals across different management scenarios.
State-and-transition models have been proposed as a framework to explicitly assess tradeoffs and win-wins for ecosystem management options (George 1992; Eastburn et al. In prep). Spider diagrams are one approach to simply illustrate relative quantities of goods and services associated with different ecosystem management options (e.g., alternative vegetation states in a state and transition model). Figure 11 demonstrates the tradeoffs and win-wins in ecosystem response based on alternative vegetation states adapted from George et al. (1992) and Huntsinger and Bartolome (1992) for the Sierra Nevada foothill gravelly-loam ecological site.
Figure 11. Spider diagrams illustrating the quantities of multiple goods and services under different ecosystem management options—resulting in alternative vegetation states (grassland (<10% canopy cover), savanna (10-49% canopy cover), and oak woodland ((<50% canopy cover)). Data on ecosystem service indicators were collected across 5,300 acres of managed oak woodland-annual grassland at the Sierra Foothill Research and Extension Center in Yuba County, California.
For each ecosystem service, the maximum distance from the center of the diagram represents the highest level of provisioning (i.e., relativized by maximum observed levels across all three states); therefore, the extent of area covered within each diagram allows for direct visual comparison of trade-offs and win-wins. For example, while the grassland state maximizes agricultural productivity, there are clear trade-offs for soil health and biodiversity and habitat relative to the other management options. The savanna state highlights a local management opportunity to balance multiple ecosystem service goals.
At the landscape scale, maintaining a heterogeneous mosaic of vegetation patches optimizes the benefits of different ecosystem management options—including increased agricultural productivity, maintaining water and nutrient cycling capacity, protecting genetic resources, and enhancing the number of habitat types. Less apparent synergies exist that cannot be directly quantified; notably, conservation of oak woodland-annual grassland landscapes has been linked to socio-economic sustainability (Huntsinger and Hopkinson 1996; Wetzel et al. 2012). Appropriate economic and social valuations for ecosystem services, taking into account tradeoffs and synergies across space and time, remain an open question (de Groot et al. 2010; Villa et al. 2014).
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List of Tables
Table 6.1. Frequency of the 20 most common annual grassland and oak woodland understory species in quadrats along 455 transects located from Mendocino and Shasta Counties to Kern and Ventura Counties (Alonso 2008).
Table 6.2. A selection of ecosystem service related studies conducted in the annual grasslands and oak woodland of California. The table is adapted from categorization of ecosystem services from Millennium Ecosystem Assessment (2005) and the Economics of Ecosystems and Biodiversity (2010).
List of Figures
Figure 6.1. Location and area of annual rangelands (oak woodlands, annual grasslands and chaparral) and other rangeland types in California.
Figure 6.2. Soft chess brome, ripgut brome and wild oats are present in most annual grassland and oak woodland ecosystems in California.
Figure 6.3. The oak-woodlands are often a mosaic of oak, grass and shrub patches.
Figure 6.4. Nitrogen cycling with major pools of nitrogen (lbs./acre) for an oak woodland-grassland ecosystem in the Schubert watershed at University of California Sierra Foothill Research and Extension Center northeast of Marysville, CA (Dahlgren et al. 2003)
Figure 6.5. Selected soil quality and fertility parameters for the 0 to 5 cm surface soils beneath an oak canopy and adjacent grasslands for three oak-woodland sites (Dahlgren et al. 2003).
Figure 6.6. Blue oak, interior live oak and coast live oak are dominant species in the oak woodlands.
Figure 6.7. Fire adapted chaparral is composed largely of evergreen, sclerophyllous shrub species that range from 1 to 4 meters (3 to 13 feet) in height.
Figure 6.8. Chamise is the most common and widespread species within the chaparral vegetation type.
Figure 6.9. Chamise resprouts from the base of the shrub following fire.
Figure 6.10. Three vegetation states representing early, middle and late in the fire cycle. The early state (PC2.1) is a savannah state with little or no shrub layer immediately following fire. The middle state (PC2.2) shows an increase in the shrub layer but low risk of crown fire and the late state (PC2.3) represents a shrub layer that has grown into contact with the tree layer increasing crown fire risk. Frequent fire tends to result in oak-woodland vegetation states dominated by an oak-annual grass community. Protection from fire and grazing results in a gradual increase in shrubs contributing to increased fuel loads and increased risk of crown fires.
Figure 6.11. Spider diagrams illustrating the quantities of multiple goods and services under different ecosystem management options—resulting in alternative vegetation states (grassland (<10% canopy cover), savanna (10-49% canopy cover), and oak woodland ((<50% canopy cover)). Data on ecosystem service indicators were collected across 5,300 acres of managed oak woodland-annual grassland at the Sierra Foothill Research and Extension Center in Yuba County, California.