The Ecological role of Cattle Grazing in Montane Meadows of the Sierra Nevada Mountains, California

PAGE 1

THE ECOLOGICAL ROLE OF CATTLE GRAZING IN M ONTANE MEADOWS OF SIERRA NEVADA MOUNTAINS, CALIFORNIA BY JENNA M. ERVIN A Thesis Submitted to the Division of Natural Sciences New College of Florida in partial fulfillment of the requirements for the degree Bachelor of Arts Under the sponsorship of Dr. Elzie McCord Jr. Sarasota, Florida May, 2009

PAGE 2

To Mom and Dad, b ecause theyre my parents and they give me pets. Julie and Bo, love you like sisters. To the mountains that I love And all the great friends I made in Fl orida, Ill be carrying you with me on the soles of my traveling shoes. ii

PAGE 3

A cknowledgements I would like to thank the Nellie May Founda tion whose generous contribution funded this research. Thank you to my thesis committee who have exhibited unparalleled patience throughout the last mont h, my thesis advisor Professo r McCord for all his advise and encouragement, the helpful staff at th e Sierra National Forest Ranger Station, the Edison International Biologists, and my friends and family who have acted as my unpaid field assistants, secretaries, navigators, tr aveling companions and bee tamers throughout this whole process. Also, thank you to my future puppy who kept me trucking. iii

PAGE 4

Table of Contents FIGURES, TABLES AND EQUATIONS ......................................................................V 1. INTRODUCTION TO SIERRAN MEADOW...........................................................1 1.1 THE SIERRA NEVADA ECOSYSTEM...........................................................................2 1.2 HISTORY OF LIVESTOCK GRAZING IN THE SIERRA NEVADAS...............................10 1.3 GRAZING IMPACTS ON MEADOW HYDRAULICS, EROSION AND SOIL COMPOSITION................................................................................................................14 1.3.1 Meadow stability and erosion ..........................................................................15 1.3.2 Run off and sedimentation in montane watersheds .......................................16 1.3.3 Soil compaction ................................................................................................18 1.3.5 Mineral redistribution .....................................................................................19 1.4 LIVESTOCK INFLUENCES ON MEADOW FAUNA COMPOSITION................................21 1.4.1 Bird Populations .............................................................................................22 1.4.2 Mammals ..........................................................................................................26 1.4.3 Fish and Amphibians ......................................................................................27 1.4.4 Insects and other arthropods ..........................................................................29 1.5 VEGETATION IMPACTS OF LIVESTOCK GRAZING.................................................31 1.5.1 Plant species composition an d Exotic species invasions ...............................32 1.5.2 Defoliation ........................................................................................................33 1.5.3 Preferential Grazing.........................................................................................................35 1.5.4 Trampling ........................................................................................................36 1.5.5 Effects on Biodiversity and Invasibility ..........................................................36 1.5.6 Shrub and Tree Expansion ..............................................................................38 2. MATERIALS AND METHODS................................................................................41 3. RESULTS....................................................................................................................49 3.1 BIODIVERSITY ....................................................................................................49 3.2 EXOTIC COVER ..................................................................................................53 4. CONCLUSIONS AND DISCUSSION......................................................................55 4.1 DISCUSSION .........................................................................................................55 4.1.1 Biodiversity ......................................................................................................55 4.1.2 Exotic Cover ....................................................................................................57 4.2 CONCLUSIONS .....................................................................................................58 APPENDIX A: SPECIES LIST......................................................................................61 REFERENCES.................................................................................................................66 iv

PAGE 5

Figures Figure 1.1 View of the Eastern Sierran slope ................................................................. 10 Figure 1.2 View of the Western Sierran slope ................................................................ 11 Figure 1.3 Relief map of California................................................................................. 12 Figure 1.4 Ely Meadow....................................................................................................15 Figure 1.5 Distribution of meadow commu nities in relation to water table ..............17 Figure 1.6 Sheep grazing in High Sierran Meadow......................................................20 Figure 2.1 Relief map of study area................................................................................43 Figure 2.2 GIS map insets for fig. 2.1 ............................................................................48 Figure 3.1 Mushroom Rock meadow .............................................................................49 Figure 3.2 Boxplot of Shannon biodiversity index values in grazed and ungrazed meadows, mushroom rock excluded.............................................................50 Figure 3.3 Boxplot of Shannon biodiversity index values in grazed and ungrazed meadows, mushroom rock included..............................................................51 Figure 3.4 Scatterplots of Shannon biodiversity index values in relationto road index, trail index and campground proximity.............................................52 Figure 3.5 Boxplot of exotic plant c over in grazed and ungrazed meadows...............53 Figure 3.6 Scatterplots of exotic plant cover in relation to road index, trail index and campground proximity...........................................................................54 Tables Table 1.1 Sierra Nevada vegetation belts by region.....................................................13 Table 2.1 Criteria for Road Classifications..................................................................46 v

PAGE 6

ECOLOGICAL ROLE OF CA TTL E GRAZING IN MONTANE MEADOWS OF THE SIERRA NEVADA, CALIFORNIA Jenna M. Ervin New College of Florida, 2008 Abstract The Californian Sierra Nevada Mount ains contain a unique ecosystem that supports a high diversity of endemic speci es. Meadows comprise only a tiny portion of the Sierras, however, their importan ce to the Sierra Nevada ecosystem far outweighs their size. Grazing has occurred in Sierras for over 100 years. Historically overgrazing has lead to detrimental changes in Sierran meadow vegetation. Today a cattle grazing in the Sierras meadows rema ins controversial. In this study, I examined the impact of grazing in 10 mont ane meadows in the Central Sierra Nevada Mountain range. Plant biodiversity and exo tic species cover was compared across grazed meadows and grazing protected meadows. Biodiversity and exotic richness among meadows was also compared to proximity of roads and recreation sites. No significant differences were found among gr azed and ungrazed meadows. Nor were recreation site and road proxim ity correlated to biodiversity or exotic ground cover. I conclude that cattle grazing, at the current stocking dens ities, do not significantly impact meadow biodiversity. Literature concerning the impacts of livestock in the vi

PAGE 7

Sierra Nevada ecosystem is reviewed and livestock im pacts on vegetation and wildlife populations are discussed. ______________________________________________ Elzie McCord Jr Division of Natural Sciences vii

PAGE 8

viii Hail to the grass, its numbers are unsurpassed -Charles Ervin

PAGE 9

1. Introduction to Sierran Meadow The montane meadows of the Sierra Nevada Mountains are economically, ecologically, and aesthetically valuable. The hu sbandry of livestock, es pecially cattle, is a practice that has played an import role in the natural, economic, and cultural history of the Californias Sierra Nevada moun tain range (Allen-Diaz 1999). The Sierra Nevada Mountains are included in the Califor nia Floristic Province, a biodiversity hotspot (Myers et al. 2000). Biodiversity hotspots are areas identified as conservation priorities due to their high endemic biodiversities coupled with large percentage of primary vegetation losses. Th ese areas contain high native plant species richness. They also support critical and often endangered plant and animal populations (Myers et al. 2000). Meadows comprise only a ti ny portion of the Sierras, however, their importance to the Sierra Nevada ecosystem far outweighs their size. Over the last century, the vast majority of meadows have served a number of commercial purposes such as seasonal grazing for domesticated livestock including ca ttle, sheep and occasionally pack animals such as horse sheep and burros (Rat liff 1985; McKelvey and Johnston 1992). Given the critical ecologi cal status of this commun ity, livestock grazing in publicly owned meadowlands and grasslands in the California Floristic Province is concerning. This debate has galvanized oppone nts and proponents, each claiming concern for the environment. While many ranchers ar gue that public grazing allotments are a low impact use and prevents land from being developed for more damaging uses (Knight 2007). Others argue that these ecologically critical areas should be protected for the 1

PAGE 10

public good and not exploited fo r commercial gain (Vavra et al. 1994, Fleischner 1994, McClaren 2 001). The arguments of each cons tituent hinge on the perceived ecological role of domesticated livestock in these comm unities. The exact nature of the interaction of livestock and their ranges is difficult to define and impossible to generalize. The impacts of livestock differ according to geogr aphical area, grazer species, past use, and myriad other ecological factors. Therefore, it is important that land managers evaluate management options in the context of the local ecosystems functions and pressures. This study seeks to elucidate the effects of livestock grazing in mid-elevation montane meadow communities. It does not seek to comment on the controversy surrounding public rangelands easements and liv estock grazing. Explicitly, it addresses only a small portion of the debated habitats and contributes to the discussion on the ecological role of domesticated grazers. This study evalua tes plant biodiversity with implications in exotic species competition in relation to livestock presence in montane meadows of the central Sierra Nevada Mount ain range. In particular, the effect of livestock grazing on biodiversity and exotic land cover will also be compared to the impact of proximity to roads, and recreation sites. Literature c oncerning the effect of livestock grazing on soil composition, watersheds, vegetation composition, and animal populations is reviewed. The role of these meadows in the surrounding landscape is also discussed. 1.1 The Sierra Nevada Ecosystem The Sierra Nevada Mountain range extends 400 kilometers from Tehachapi Pass to the south, to Lake Alamar and the beginnings of the Cascade Mountains in the north 2

PAGE 11

(Fig. 1.3). T his range has been described as a great granite block, tipped on edge such that the western slope (Fig. 1.2) meets the Gr eat Central Valley at a gradual 2% to 6% grade while the eastern slope (F ig. 1.1) rises quite sharply fr om the mid-elevation valleys and plateaus that mark the beginning of the great basin. A wide variety of climatic conditions prevail throughout due to the comple x topography of the Sierras, coupled with Figure 1.1 View of the Eastern Sl ope of the Sierras from Benton Cr ossing campground (Photo taken by Diane Ervin September 2008). Note the steep slope of the peak s in the background and sparse vegetation characteristic of the Eastern Sierras. their geographic position in re lation to the San Joaquin, Costal Ranges and the Pacific Ocean. One may encounter climatic conditi ons ranging from those resembling those 3

PAGE 12

nearer to the Mexican border to conditi ons nearly resembling the American Artic Regions.In general, the Sierras experience dr y, cool summers and m ild winters. Ninty five percent of the precipitation falls in the winter between months of October and May. Summer high temperatures range from 80 to 100 F and winter lows range between 0 and 30 F depending on elevation. The westward sl ope (Fig 1.2) experiences an average of 75 inches of rain fall per year. The temp erature is reduced approximately 1 F and precipitation increases 2-4 inches for each 300 ft increase in elevation. The east slope (Fig 1.1), in contrast, experiences a distinct rain shadow effect re sulting in near-desert conditions. It experiences an average of 20 inches of precipitation per year. Some areas, Figure 1.2 Tuolome pass (Photo by Diane Ervin September 2008). The western slope of the Sierras is more moist and lush than the eastern slope, vegitation is characterized by mixed conifer forest in the higher elevations and blue oak savanna/chaparral in the lower elevations 4

PAGE 13

Figure 1.3 Relief map of California's bioregions. The Sierra Nevada Bioregion as specified by Jepson is outlined in light blue. Research for this stud y was conducted in the area indicated by the black arrow (Department of Fish and Game Website 2005). 5

PAGE 14

such as the alpine regions and the dry, spar se eastern slopes, e xperience more drastic clim atic swings (Storer 2004). Every rise in elevation experiences its own sequence of seasonal changes, weather, and biological patterns. For this re ason, plant and animal distributions are usually disc ussed in relation to a series of belts that roughly follow elevation contours along the leng th of the Sierra Nevadas. The boundaries of these belts are variable as extreme topographic bands are higher on the east slope than the west slope. Also bands extend higher on warm s outh facing slopes and te nd to be lower on north facing slopes and shaded gullies. The bands and their approximate elevation ranges from lowest in elevation to highest are listed in table 1.1. Table 1.1 : Sierra Nevada vegetation belts elevation by region elevation of vegetation belts by region in feet above sea level Belt name Southern Central Northern Foothill 1 1250-5000 800-4000 500-3000 Lower Montane 1 4 3600-6300 4000-6500 900-5500 Upper Montane 1 8000-10000 6500-8000 550-7500 Sagebrush 2 n/a 6000-7000 4200-5600 Sub-alpine 3 8700-1100 7200-9150 7200-9150 Alpine 9000-1100 n/a n/a 1 western slope only 2eastern slope only 3 area commonly referred to as the High Sierra 4 also referred to as the mixe d conifer or yellow pine belt data from Storer et al. 2004 The meadows observed in this thesis ar e located in the uppe r and lower montane belts, therefore, these belts will be examined in more detail. The lower limit of the lower montane belt is dominated by mixed conifer woodland and ponderosa pine ( Pinus ponderosa L .). In the upper regions, white fir ( AIbes concolor (Gord. and Glend.) Lindl. ex Hildebr.), sugar pine ( Pinus lambertiana Dougl.) and jeffery pine (Pinus jeffreyi Grev. and Balf.) are equally represented in the mi xed conifer woodlands. Also widespread in 6

PAGE 15

this belt is the m ountain chaparral habitat. This habitat is dominated by low brush species including manzanita ( Arctostaphylos sp Andans), white thorn ( Ceanthus cordulatus Kellog), bitter cherry ( Prunus emarginata (Douglas. ex Hook)), bush chinquapin ( Castanopsis sempervirens (Kellogg) Dudley ex Merriam) and huckleberry oak ( Quercus vacciniifolia Kellogg). Summers, at this elevation belt, are warm with average high temperatures ranging between 80 and 93 F, and winters are cool with average minimum temperatures of 22-34 F. The growing season is four to seven months long and transpires over the warm summer m onths. Some winter snowfall occurs. The lower montane belt experiences the heaviest l ogging activity in the Sierras. The Upper Montane belt includes mixed conifer woodl and with slightly different species composition: white fir, je ffery pine, quaking aspen (Populus tremuloides Michx.) and lodge-pole pine (also commonly re ferred to as tamarack pine) ( Pinus contorta ver. murayana (Grev. and Balf.) Engelm.). These tree s are often covered in a layer of wolf lichen ( Letharia vulpine (L.) Hue). Winter snows are often heavy and snow cover is persistent. The growing season is much shorte r, three to four months. Biologists often lump the two montane regions together when discussing plant and animal distributions (Storer et al. 2004). Montane meadows are defined as wetlands or semi-wetlands supporting a cover of emergent hydrophytes and mesophytes (Fig 1.4 ). In contrast, meadows of the alpine and subalpine regions are dry and dominated by shrubs. Montane meadows in the Sierra Nevada provide a large number of products and services in the ecosystem, although they comprise only 10 percent of the land area. Th ey provide the bulk of forage in grazing allotments and are necessary for the producti on livestock products and the raising of 7

PAGE 16

recreational stock in the mount ainous areas. They are necess ary resources and habitats for a variety of wildlife populations includi ng native large grazers such as mule deer ( Odocoileus hemionus Rafinesque). Timbered edges be tween meadows and forests create ecotones, transitional areas that harbor a high biodiversity of plants and wildlife (Ratliff Figure 1.4 Ely Meadow (Photo by Diane Ervin October 2008) 1985). Also they provide recreational and economic services as scenic vistas and favored camping spots for forest, park and wi lderness visitors (Ratliff 1985; Rundel et al. 1977). Meadows are found in nearly every fore st type throughout the montane belts. They can range from a few square meters to large, several-hundred hectare swaths of treeless land. The orig ins of these meadows are variab le. Favorable conditions must prevail for the maintenance of mea dow-type vegetation at any one site: 8

PAGE 17

a relatively rocky im pervious bedrock floor, a upper drainage area of sufficient si ze to maintain a high water table in the meadow site well in to the growing season, a gentle gradient, and, a favorable drainage area to slope ra tio (large drainages with steep slopes generally do not contain meadows) (see section 3). Many meadows result from the infilling of gl acial lakes or valleys. Mid-elevation (1500-2500 ft) meadows can also be associated with stream valleys that have never been glaciated (Wood 1972). Meadows can be the product of logging if substrate is appropriate and drainage is poor. Meadows may also be pr oduced by the infilling of ponds with peaty materials due to the growth of vegetative materials inside them. Soil analysis and the presence of mature tree stumps beneath the soil of some meadows suggest that these sites have unde rgone regressive succession (Rundel et al. 1997). Over geologic time, some meadows have repeated ly developed, disappeared, and redeveloped in the same geographic locations (Ratliff 1985). Within a single meadow, plant species are distributed in zones according to water table depths and soil texture gradients (Fig 1.5). The commu nity associated with the highest water table is characterized by an aerobic soil conditions and is dominated by sedges ( Carex sp.) and exhibits low species rich ness and biodiversity (Dwire et al. 2006 ). In my study, communities in this zone to be comprised of dense and nearly monotypic stands of Carex nebraskensis Dewey and Carax vesicaria L The zone of intermediate species diversity and moderately high water table is termed the moist meadow community (Dwire et al. 2006 ). A mixture of sedges, reeds ( Juncus sp. ), grasses 9

PAGE 18

( Poacea sp.) perennial forbs and m osses characteri ze this community. The area with the lowest water table is termed the dry meadow community. It is characterized by aerobic soils, high biodiversity and species richness. The composition of plants in this zone is similar to that of the moist meadow commun ity. The differences between the moist and dry zone plant communities lie in a greater presence and cover of hydrophytic species in the moist community and a greater presence of mesophytes in the dry community (Dwire et al. 2006). Figure 1.5 Distributions of meadow communities in relation to water table height (Data from Dwire et al. 2003, illustration Jenna Ervin 2009) 1.2 History of Livestock Grazing in the Sierra Nevadas The pool of knowledge concerning the c ondition of California rangelands before 1860 is based primarily on the accounts of early European explorers of the region. From these accounts, it is suspected that rangelands at the time of early European exploration 10

PAGE 19

were fertile and dom inated by perennial bunchg rasses (Kinny 1996). Prior to European colonization and the introduction of do mesticated livestock, burning by Native Americans modified grasslands, meadows a nd woodlands. This practice insured open defendable territories, and cleared grasslands for game animals (Anderson and Moratto 1996). Cattle ranching became the first industry in California with the establishment of missions along the southern California coas t (Ratliff 1985). Grass species found in adobe bricks of early missions showed evidence that several species of non-native annual grasses had already become established prio r to Spanish settlement (Hendry 1931). The first introduction of cattle wa s 200 head at San Diego in 1769 (Kinney 1996). The first flock of sheep, was brought into San Die go in 1770. Sheep and cattle production was concentrated around large mission-owned ranches and individual holdings until 1850. At this time feral herds of cattle roamed the state while sheep populat ions were kept at a minimum due to predation and high production costs (Kinney 1996). The discovery of gold in California in 1849 marked a point of unprecedented change. Settlers and prospectors rushed to California. This boom in population and industry coincided with a sharp increase in the demand for lumber, meat and land. The focus of livestock production turned away from leather, wool and tallow as demand for meat increased. Cattle and sheep numbers gr ew dramatically to meet demand (Ratliff 1985). There were 250,000 cattle recorded in California in 1850, however that number climbed to 1 million by 1860. That number trip led in two additional years. Sheep were also imported into the state in equally da unting numbers (Ratliff 1985). At this same time logging activity increased to meet the dema nds of new mines and settlements. Strip 11

PAGE 20

m ining, an ecologically destruc tive practice, came into use in the portion of the western Sierras known as the Mother Lode. This pr actice involved the use of high-powered water jets to blast away soil and vegetation, exposing veins of gold-rich quartz below (Storer 2004). Very few accounts of montane meadow s existed before 1860. Most accounts of early explorers focused on the foothill vegetation. More information was revealed only after extensive surveys, circa 1860, and the ac counts of John Muir in 1869. Muir (1989), who traveled the area extensiv ely in the 1860s and 1870s, provid ed some descriptions of the meadow vegetation he encountered. Thes e accounts generally de scribed meadows as lush, populated by chest high vegetation, comm on forbs and brush species do not differ significantly from meadows as they appear to day. The practice of dr iving herds into the mountains during the summer began in respons e to series of droughts and floods in 1863 and 1864 (Kinny 1996). These natural pressures as well as the steady conversion of lowland rangelands into farmlands prompted ranchers to move their herds seasonally in search or forage. It is duri ng this period that the most si gnificant damage to the Sierran rangelands occurred (Kinny1996, Du ll 1999). Most damage is attributed to sheep grazing. The status of the Sierra Nevadas as comm on land allowed the unregulated movement of thousands of head over the passes (fig 1.6). In addition to the defoliation and trampling by flocks, herders routinely burned huge tracks of land to prevent the growth of brush. 12

PAGE 21

Figure 1.6 Sheep grazing in north side of Crabtree Meadow (elv. 10,400 feet), Sequoia National Park, Sierra Nevada, California, 29 July 1895 (V ankat 1970) The frequency and intensity of these distur bances allowed for little or no nutrient recovery in the soil, destruction of potentially regenerative root systems, and erosion of stream banks. It is hypothesized that lasting and extensive damage was done to meadow ecosystems. However, the extent this damage is hard to estimate without more detailed descriptions of pre-grazing condi tions (Kinney 1996; Dull 1999). In 1881, due to the efforts of naturalists and concerned citizens such as John Muir, Yosemite and the central Sierra giant sequoi a groves were granted to the State of California. This was quickly followed by the es tablishment of the Sierra Forest Reserve. While the establishment of public lands in the Sierra Nevadas was a boon for 13

PAGE 22

Sierra Nev ada conservation, degradation contin ued due to a lack of effective, year-round enforcement personnel and inadequate violation penalties (Kinny 1996). Effective control was gained in 1905 with the establishment of the Department of Agriculture and the instatement of grazing pe rmit requirements. Today, sheep have been replaced completely by cattle in the Sierra n rangelands. Grazing is prohibited inside National Parks with the exception of a few sm all sections still unde r partial control of private owners (McKelvey a nd Johnston 1992; Kinney 1996). Cattle production in the Sierras has d ecreased since the 1920s. However, the presence of pack animals and recreational backpacking and camping has significantly increased over this period (R atliff 1985). A surge of envir onmental concern in the 1970s marked a period of renewed interest in the Sierra Nevada meadow ecosystems. At that time several key research programs were establ ished with the sponsor ship of the Sierra Club and the National Park Service in associ ation with the University of California (Kinney 1996; Storer et al. 2004). Based on the change of focus in recent research, it appears that interest in mea dow ecosystems has since declin ed; however, the pressures of multiple uses and the controversy of grazing still persist today. 1.3 Grazing impacts on Meadow Hydraulics, Erosion and Soil Composition Livestock-driven change to meadow ve getation and alterations to wildlife populations, often result from changes to abiotic meadow components (i.e. meadow hydraulics, erosion, and soil composition). The mo st important factor in the formation of a meadow is the presence of a high wate r table (Wood 1975). It follows that montane meadows are often associated with watersheds and riparian systems. This chapter seeks 14

PAGE 23

to illustrate the impact of livestock on the abiotic components of th e meadow ecosystem. Specifically, the role of defoliation, tram pling, and nutrient redistribution will be examined in relation to soil, water and sedimentation. Thes e processes will be discussed in the context of meadow stability and function. 1.3.1 Meadow stability and erosion Meadow stability can be considered in terms of biological and geological stability. Biological stability refers to the pe rsistence of meadow species. Alternatively, geological stability refers to the persistence of the geological conditions which provide an environment favorable for meadow formati on and maintenance (Benedict 1982; Allen 1987; Tillman and Downing 1996). Anthropogeni c activities (i.e. livestock gazing) contribute to both biological and geological instability of montane meadows (Wood 1975; Ratliff 1985). A geomorphic instability threshold is the point at which meadow topography can no longer support meadow vegetation and mead ow succession into forest commences. Meadow topography approaches a geomorphic instability threshol d as the sides of meadow become steeper and the deposition of sediment into the meadow center increases (Schumm 1977). A natural sod of sedges, gr asses, mosses, and forbs protect meadow soils from sheet erosion, or the removal of soil from continuous areas (Ratliff 1985). Livestock trampling breaks up this sod. Th is damage may not be repaired for many subsequent years (see section 1.4). Breakup of sod is the major driver of meadow instability (Wood 1987). Erosion occurs when a meadow becomes geologically unstable. Erosion removes protective sod, results in sh eet erosion of productive topsoil, and drives 15

PAGE 24

gully formation. Gully formation, the cutt ing of deep strait channels through the meadow, alters meadow hydrology, lowering th e water table and decreasing meadow moisture. Meadow moisture is an impor tant factor in determining vegetation composition, exotic invasibility and reactions to grazing (s ee section 1.4). Usually the changes are undesirable in nature. It is im portant to note however, that ungrazed intact sod may not protect a meadow from gully er osion if the meadow is already near its geomorphic stability thres hold (Wood 1975; Ratliff 1985). 1.3.2 Run off and sedimentation in montane watersheds Meadow vegetation plays an important ro le in controlling runoff and sediment movement into associated waterways (Cla ry 1987). Meadow vegetation lies between sediment sources and streams, reducing suspended sediment from overland runoff, providing clean water for human and animal c onsumption and favorable habitats for fish and amphibians. These vegetation buffers are re ferred to as Vegetation Filter Strips (VFS) (Fraiser et al. 2002). It was once widely accepted that water moved through a meadow as sheet flow. More recent research, however, suggests that the majority of water movement occurs in the form of microchannels; small flow pa ths through vegetation. These flow paths are defined as measuring about 10-20 cm wide and 1-2 cm deep (Fraiser et al. 2002). Limited sheet flow does occur over short distances between microchannels but does not account for a large portion of water movement. Mo re sinuous microchannels within the VFS retain greater overland sediment loads. Effici ency of VFSs, therefore, depends on surface microtopography, vegetation cover density and type, and buffer stri p slope and length. 16

PAGE 25

Increased stem density and stem diameter are the greatest contributing factor to microchannel sinuosity (Fraiser et al. 2002). Flanniken and colleagues (2001) found th at simulated grazing significantly decreased water infilt ration, increased sediment flow, and runoff through alterations in soil compaction, defoliation, and physical damage to vegetation. Grazed plots exhibit less above ground plant biomass and decreases in st em density. Decrease in stem density is mainly due to trampling by hooves and not to trimming (Flanniken et al. 2001). Higher stem density decreases sediment flow by forc ing water into a slower more sinuous path. Straightening of micro-channels can lead to greater gully-like act ion of microchannels characterized by faster-flowing water and steeper channel banks. Gully-like action has a greater potential for erosive forces and sediment movement (Flanniken et al. 2001) Grazing defoliation, therefore, can impede the ability of meadow vegetation to function as a VFS to adjacent watersheds. Increas es in sediment movement were most exacerbated by grazing in the first meter of vegetation surrounding a riparian system (Fraiser et al. 2002). Vegetation height is a less im portant contributing factor than vegetation density (Fraiser et al. 2002). Trampling associated with cattle grazi ng also decreased th e amount of water stored in small depressions in the path of water flows. While this decreases water infiltration into the soil, it also causes flow events to be more uniform, causing water to move into streams at a steady rate rather th an in one large surge. Uniform water flow regimes tend to increase the effectiveness of riparian vegetation buffers (Flanniken et al. 2001). Fraiser and colleagues (2002) suggested that properly managed cattle grazing may 17

PAGE 26

be compatible with management goals gi ven that grazing may stabilize water flow regimes through meadows. 1.3.3 Soil compaction Prolonged trampling by large herbivores may result in changes in soil pH, soil nutrient compositions, and soil nutrient distri bution. Trampling of soil by livestock has two major consequences apart from the disr uption of above ground vegetation. The first, cutting of meadow sod and subs equent erosion, was discussed extensively in the previous section. The second is soil compaction. Co mpacted soil resists root growth and decreases water infilt ration. Compaction alte rs the soil surface structure, creating favorable conditions for erosion. The degree of soil compaction is measured by changes to bulk density and non-capillary pore space. Increasing compaction is directly related to bulk density and inversely related to non-cap illary pore space (Lull 1959) Comparisons of soil bulk densities are only useful in meadows of similar soil moisture contents, as soil moisture tends to increase soil mass but not necessarily mineral de nsity (Laycock and Conrad 1967). It has also been shown that compaction around high traffic areas could decrease soil pH as much as one pH unit, however the mechanism by which this occurs is unclear (Leonard 1968, Lockaby and Dunn 1978). A study of simulated shearing and trampling found only slight increases in stream bank compaction in response to mid-level (1 AUM (animal unit month) ha-1) defoliation and trampling. Grazing was simulated by c lipping vegetation hei ght, applying random hoof-imitator impacts, and depositing urin e and fresh manure. Heavy season-long simulated grazing (~9.5 AUM) activiti es resulted in significant whole stream changes. 18

PAGE 27

These chang es include exacerbating factors th at decreased stream bank stability and increased erosion such as lowered root biom ass and increased stream bank slope (Clary and Kinny 2002). Melinda and colleagues (2002) conducted a soil survey in which cattle were grazed in a meadow site which had previous ly been grazing-excluded for 45 years. Grazed meadows exhibited significantly increa sed bulk soil density and decreased water infiltration rates compared to the pre-disturbance levels. Meadows soils with high levels of organic content resisted the effects of trampling but mineral soils beneath this still experienced compaction. Compaction effects we re most prominent at soil depths of 5-10 cm of soil The study site soil recovered to pre-disturbance levels in one year, possibly due to the high amount of freeze-thaw events and high organic content that are characteristic to montane meadows. The degree to which trampling damages stream banks and sod, and influences compaction, is a function of soil moisture. We t soils are more yielding than dry soils; therefore wet soils are impacted to a greater degree than dry soils (Clary and Kinney 2002). Clary and Kinney (2002) recommend th at cattle be excluded from meadows during the wettest part of the year. A gene ral suggested guideline is to allow grazing only when soil moistures are less than 10%. 1.3.5 Mineral redistribution The majority of the nutrients consumed by meadow grazers are returned in the form of urine and excreta. Therefore, little mineral nutrients are lost to domesticated grazers. However, the presence of any grazer will result in the redistribution of nutrients 19

PAGE 28

from preferred browsing areas to preferred lo afing areas, spaces where cattle congregate but do not feed (Ratliff 1985). Blank and colleagues (2006) found that differences in soil nutrient composition were generally more pronounced around meadow edges. They attributed altered patterns of nutrient distribution to five grazing-related processes; landscape redistribution of elements due to urine and feces deposition, compensatory plant growth, changes in soil microbial community, soil compaction and li tter accumulation. Even though nutrient changes are observed at meadow edges, cattle tended to spend a majority of their time mid-meadow or streamside while forest-m eadow edges received a greater portion of loafing use. Ratliff (1985) posits that cattle tend to congregate in the dryer portions of meadows early in the season when vegetation is lush. but will move to the wetter meadow portions later in the season as meadow aridity increases. Local mineral redistribution impacts vegetation composition by creating favorable condition for plants adapted to the altered mineral state. Another important vector of mineral re distribution is the deposition of human dung into the meadow ecosystem by recreationa l hikers (Reeves 1979). Redistribution of nutrients by humans is different from that of livestock because these nutrient loads are being redistributed from a source outside the meadow rather than within the meadow. Because hikers tend to use latrines inst ead of randomly distributing dung about a meadow, nutrient redistributi on can be ultra-concentrated to only a few areas (Ratliff 1985). 20

PAGE 29

1.4 Livestock influences on meadow fauna composition Livestock can influence the composition of meadow fauna both through indirect and direct pathways. Direct influences include competition for forage and space. More often, cattle and other grazing live stock exert indirect influenc es by altering the structure of the vegetation communities (Wieren 1998) and soil and stream bank morphologies (Knapp et al. 1998). It has been argued that domesticated ungulates may act as keystone species in some grasslands by filling the ecological niches of absent or extinct large grazers. An example of such an interaction would be cat tle filling the role of American Bison on the Great Plains (Knapp et al. 1999). However this concept can only be applied in systems where large grazers have been present in the past and onl y at low stocking densities (Fleischer 1994). Even given these parameters, it is controvers ial if domesticated grazers, which usually exhibit different su ite of habits, truly can replace absent megafauna. In systems where cattle do not act as ke ystone species their effect on vegetation is dependant on livestock density. At low to medium densities cattle exert little effect on vegetation communities. Cattle, present in high densities, in systems where large grazers were not part of the ecosystem, can result in dramatic negative impacts (Van Weiren 1998). Kinney (1996) posits that because large grazers where present in prehistoric times, domesticated livestock grazing can provide vital disturbance event in montane meadows, comparable to the importance of fire dist urbance. All montane meadows in this study were at one time grazed intensively by sheep (Personal communication, rangeland specialist 2008). Hence, it is difficult to ascertain what the natural state of the 21

PAGE 30

ecosystem truly was and whether livestock graz ing shifts the ecosystem toward that state (Kinney 1996). By examining areas of cattle exclusion, it may be possible to understand the way in which livestock presence influences meadow ecosystems in a comparative context. Livestock tends to benefits some species and exclude others (Fleischer 1994). These interactions are highly species speci fic and montane meadows are by definition highly diverse (Ratliff 1985). Therefore, an exhaustive review of influences on all meadow species would be long indeed. In this chapter, I will atte mpt to address those species that are of greatest in terest to management efforts. These species include those that are threatened, endangered, or of econo mic significance. However, I would like to stress that this list in no way completely covers the complex web of interactions that occur in montane meadows and that may be impacted by cattle presence. 1.4.1 Bird Populations Montane meadows play a unique and crucia l role in the life hi story and ecology of several groups of Sierran birds. Two sp ecies currently listed under the endangered species act, the Willow Fly Catcher ( Empidonax traillii Audubon) and the Gray Owl ( Strix nebulosa Forster), nest almost exclusively in meadows or rely heavily on montane meadow habitat. Other groups that depe ndently or heavily utilize montane meadows include those that require healthy meadows for successful breeding, southern migrants, up-slope dispersers (montane breeders th at aggregate in meadows during the postbreeding period), and forest/edge occurring species (Wilkerson and Seigel 2001). 22

PAGE 31

Birds do not generally respond to the presence of livestock but rather to changes in vegetation due to grazi ng (Bock and W ebb 1984; Saab et al. 1984; Fleischer 1994). These impacts include shortening, shearing a nd trampling of the vegetation as well as changes in vegetative community structures in response to alteration of soil and water dynamics. Heavy grazing can create favorable habitats for some birds and threaten populations of others. Commensalisms occur between meadow dwelli ng bird species that follow cattle because they stir up pr ey. These include cattle egrets ( Bubulcus ibis L.), brown headed cowbirds ( Molothrusa ater Boddaert), and starlings (Sturnus vulgaris Rafinesque). Dead cattle potentially increas e the incidence of carrion seeking birds. Domesticated grazers tend to decrease the a bundance of birds that nest in tall grass by reducing suitable nesting habitat. On the ot her hand, cattle can increase habitat for water birds by creating more open water space in reed beds (Van Wieren et al. 1996). Brown-headed cowbirds associate with large ungulates. In Sierran meadows, cowbird presence increases in conjunction wi th cattle grazing. Brown headed cowbirds are well known nest parasites of many breeding neotropical landbirds (Airola 1986, Verner and Ritter 1983). Before settlement of the west, cowbirds where associated with the giant bison herds of the Great Plains region. Cowbirds expanded their range with European expansion, the clearing of forest land, and the establ ishment of livestock ranges (Rothstein et al 1980). The range of cowbirds now encompasses the entire United States. They are sufficiently numerous in montan e meadows to threaten sensitive populations which they parasitize (Rothste in et al 1980; Saab et al 1984). Brown headed cowbirds, therefore, are the medium in some indirect livestock impacts on other avian species. Cowbird parasitism may play a key role in the decline of willow flycatchers 23

PAGE 32

( Empidonax traillii), a threatened passerine bird whose few rem aining populations inhabit isolated montane meadows of the Sierra Nevadas (Sedgwick and Knoff 1988; Sanders and Flett 1989; Sedgwick and Iko 1999,) Willow flycatcher disappeared from the majority of their former range in Ca lifornia (Sanders and Flett 1989). Loss of meadow habitat to reservoir filling, hydroelectric devel opment, and tamarack pine encroachment has contributed to the declin e of willow flycatcher populations (Serena 1982). Cowbird parasitism also appears to be a significant contributo r to nesting failures in lowland willow flycatcher populations. Howe ver, there are less recorded incidences of cowbird parasitism in higher elevations (Stafford and Valentine 1985; Flett and Sanders 1989). Flett and Sanders (1989) suggested that peak cowbird breeding season often occurs before Willow flycatcher egg-laying in high elevations. This lack of reproductive overlap may be responsible for the decrease d incidence of high elevation parasitism. Direct damage to willow flycatcher nests, nesting areas, and individuals by livestock also contributes to flycatcher dec line. Willow flycatchers tend to place their nest near the edge of willow clumps and al ong livestock trails, maki ng them particularly vulnerable to livestock disturbance (Flett and Sanders 1987). When cattle pass through willow thickets, they can knock over nests of willow flycatchers and other passerines. The potential for cattle to dist urb nesting sites depends on th e overlap of nesting season and cattle grazing season. Most bird species that nest in the meadows, flycatchers included, incubate eggs or nestlings around late June (S tafford and Valentine 1985). These species are vulnerable to nest upset from late June till early July, a time when cattle are just being moved in to the meadows (Ratliff 1985) 24

PAGE 33

In addition to willow flycatchers, Sande rs and Flett (1987) have identified 16 other birds species breeding in montane mea dows that could be directly impacted by cattle presence. Willow-nesting species include Yellow and Wilson's Warbler ( Dendroica petechia L and Wilsonia pusilla Wilson), White-crowned Sparrow ( Zonotrichia leucophrys Forster), Song Sparrow ( Melospiza melodia Wilson), and Redwinged Blackbirds (Agelaius phoeniceus L.). Ground nesting birds are part icularly vulnerable to livestock trampling. These species include Canada Goose ( Branca canadensis L.), Mallard ( Anas platyrhynchos L.), Cinnamon Teal ( A. cyanoptera Vieillot), Virginia Rail ( Rallus limicola Vieillot), Sora ( Porzana carolina L.), Killdeer (Charadrius vociferus L.), Spotted Sandpiper ( Actitis macularia L.), Common Snipe ( Gallinago gallinago L.), Wilson's Phalarope, ( Phalaropus tricolor Vieillot), Savannah Sparrow ( Passerculus sandwichensis Bonaparte), and Lincoln's Sparrow ( Melospiza lincolnii Audubon). Conversely, grazing increases the relative amount of bare ground. Thus, habitat and abundance of ground nesting birds is increased. These species include mourning doves, quail meadow lark (Van Wieren 1998). In a review of nine studies of riparian habitats including willow thickets, Saab and colleagues (1995) found consistent responses of riparian avifauna to livestock grazing treatments. Of the species listed above, two, red winged blackbirds and willow flycatchers, consistently showed populati on density decreases in grazed vs. ungrazed plots. Others showed inconsistent responses to grazing among meadows. 25

PAGE 34

1.4.2 Mammals Domestic livestock interactions with mammals can be divided into interactions among herbivores and those among carnivores Among herbivores, cattle grazing can have a positive influence on small, grass eat ing herbivores such as ground squirrels and jackrabbits. However some rodents, such as mice species, preferred ungrazed meadows (Page et al 1978). The pika ( Ochotona princeps Richarson), a threatened r odent of the High Sierras, may be negatively impacted by cattle grazing. Pikas occur only in isolated populations on high altitude peaks throughout the West. Th ese populations have suffered expatriation from their ranges over the last several decad es. A significant correlation has been found between pika extirpations and grazing of livestock in and around pika habitats. Sarr and colleagues (2003) suggest that the tramping of soil and vegetation w ithin 20-50 m of the talus, loose broken rock in wh ich pikas nest, decreases forage and causes pikas to travel further in search of food. Population impact s are due more to increased predation risk rather than any exploitiv e competition for food (Sarr et al. 2003). The introduction of cattle and sheep onto the rangelands of the Sierra Nevadas had a profound impact on the native mule deer populations. Sheep compete heavily with native mule deer by grazing Sagebrush (Artemisia sp .). Sheep grazing is uncommon on the western Sierra slopes but is still relativ ely common on the more arid eastern slope. Unlike sheep, cattle and deer share relatively little dietary overlap. Deer consume mostly woody browse (i.e. the soft shoots and leaves of shrubs and trees) and cattle and sheep eat mostly graminoids. However, dietary overlap in creases as availability of preferred forage decreases (Beck and Peek 2005). 26

PAGE 35

Detrim ental effects from cattle grazing appear to arise from interference and effects on vegetation cover. Mule deer prefer to forage in grazed pastures rather than ungrazed pastures when cattle are absent (Rao tzie and Bailey 1991). However, mule deer are cryptic animals and tend to avoid cattle wh en they are present. Moderate to heavy grazing decreases the amount of vegetation cove r in riparian and as pen habitats (Loft et al. 1987). In a radio-collar study, Loft and colleagues (1993) observed that 27 female deer, tended to avoid cattle when they are moving through their home ranges. Also, heavy grazing triggered habitat sh ifts of to less preferable forage in the late summer as competition for preferable habitat increased. Kie and colleges (1991) found that deer spent more time feeding and less time resti ng with increased cattle stocking rates. Mountain lion ( Puma concolor L.) home range size depend on mule deer distribution (Grigione et al. 2002). Redistribution of mule d eer due to cattle avoidance or attraction to a formally grazed site may in fluence mountain lion di stribution. However a causal correlation betwee n cattle grazing and mountain lion home range size has yet to be determined. 1.4.3 Fish and Amphibians Most montane meadows, and all meadows surveyed in this st udy, are associated with several types of watershed systems. Some are associated with riparian systems, while others are associated with mid-successional streams or lakes. Therefore, water dynamics in these meadows have important c onsequences for aquatic biota and larger bodies of water, such as lakes and rivers, a ssociated with them. In meadows that are 27

PAGE 36

associated w ith a mid-sized stream or pond, grazing can significantly impact fish populations. Golden trout ( Oncorhynchus mykiss aguabonita Jordon) are native to the Sierra Nevadas and are endemic to only two stream drainages there (Knapp et al. 1998). Golden trout populations have declined dr amatically in the la st 100 years due to competition, hybridization and predation by stoc ked, non-native, trout species such as the rainbow trout. Golden trout numbers have b ecome so low that they are currently being considered for listing under th e endangered species act (Fish and Wildlife Website 2009). Cattle presence in stream beds often results in destabilization of stream banks, leading to the formation of wider shallower stream be ds. This process actually appears to encourage the frequency of golden trout, as wide shallow stream beds are their preferred spawning habitat for golden trout (Knapp and Vrendburg 1996). Amphibian populations have declined drasti cally in the later half of the last century (Drost and Fellers 1996) It is suspected that livestock grazing in montane meadows and riparian areas may be one of several contributing factors to this decline. Of the Sierran amphibians, true frogs and toad s have undergone the most intense declines (Serman and Morton 1993; Drost and Fellers 1996). These species include the Yosemite toad ( Bufo canorus Camp), and the yellow-legged ( Rana muscosa Camp) and red legged ( Rana aurora draytonii Baird & Girard) frogs, both of which are endangered. Yosemite toads are associated with th e vegetation of high montane an d subalpine meadows. They are primarily active during the spring summer and early fall when meadows are stocked with cattle. The toads prefe rred sedge and rush dominated breeding areas found in the warm, slow-moving waters of open meadows. Because of the overlap in habitat and 28

PAGE 37

occurrence, cattle disturbance has been identifi ed as a potential contributor to the recent decline of Yosemite toads. In addition to trampling cr yptic juveniles and larvae (Jennings 1996) and polluting breeding ponds, ca ttle also have the potential to inflict indirect damage by changing water regime water quality, and meadow microtopology. Also, removal of vegetation can increase pred ation, increase risk of desiccation and remove important food resources (Jennings and Hayes 1994). However, the decline of amphibian species in the Sierras is most likely a result of several cumulative and long term causes and removal of cattle grazing alone may not impact the trend. The connection of cattle presence to amphibian decline is supported mostly by anecdotal observation. Currently the USDA Forest Se rvice is taking measures to quantify and minimize the potential impact of cattle on threatened amphibian populations. These efforts include long term monitoring progr ams as well as the closing of several previously utilized grazing permits in 2001 (USDA Forest Service Website 2009). The consequences of these efforts are yet to be evaluated. 1.4.4 Insects and other arthropods Management of insect populations has not been a major issue in the Sierras and there has been little focus on quantifying their populations. As with other animal taxa, cattle tend to attract some insects and exclude others. Insect species that are generally associated with livestock will potentially incr ease their populations in response to cattle presence. Parasites, such as bot flies and warb le flies are specific to cattle while ticks, lice, and horse flies favor ungulates in genera l. No studies have quantified how cattleborn generalist parasites might influence na tive ungulates such as mule deer. Dead 29

PAGE 38

anim als and dung can attract carrion seekers and dung beetles (Van Weiren 1998). Many insects require tall herbaceous vegetation to thrive and are seriously affected by changes in habitat structure. For example, speci alist Sierran lepidopte ron populations depend on abundant coverage of obligat e wetland plants which can be impacted by cattle presence (Strathmann 2005) Factors that have the larges t negative impact on inverteb rate populations in grazed systems are fewer individual flowering pl ants, less litter, reduced standing crop, disturbance by treading, less prey for predators and more stream microclimate. A study of landscape factors shaping bumble bee popul ations in montane meadows found that cattle and sheep grazing decreased species richness and abundance of meadow bee communities (Adams 1975). Sheep tend to remove nearly all available floral resources and create vegetation gaps (Anderson and Ca lov 1996). Therefore, th e effects of sheep grazing on pollinators could be more important than the effects of cattle grazing Out of the wealth of insects present in montane meadows, grasshoppers appeared to be one of the most researched. Grass hoppers play an important economic role in grasslands. In 1983 it was estimated that th e yearly average estimated loss of cattle forage to grasshopper grazing was 13.35 milli on metric tons for the 262 million ha of western rangeland (Hewitt and Onsanger 1983). Some grasshopper species benefit from heavy grazing (Fleischner 1994). These spec ies require both short vegetation for egg deposition and high dense vege tation in the nymphal and imaginal stages. Variable vegetation heights due to grazing can provide these conditions (Wierren 1998). 30

PAGE 39

1.5 Vegetation Impacts of Livestock Graz ing Grazing produces two distinct classes of vegetation change in meadows. The first is the immediate cosmetic damage as a resu lt of trampling and shearing. However, the importance of this change is debatable and hi nges on the intended purpose of the land in question, namely, recreation or livestock pr oduction (Vale 1987). The second class of changes is those that impact the growth of individual plants and communities as a whole and carry over through multiple growing season s. Domestic grazing is associated with several common vegetation level processes that can adversely affect meadows, some of which have already been discussed. These pr ocesses include exo tic species invasions, defoliation, selected species exclusion through preferential grazing, trampling, redistribution of nutrients within meadows, am plification of detrimental rodent activities, woody plant encroachment, and erosion associat ed with vegetal cover loss (Ratliff 1985). The effect of different intensities of domestic grazing on montane meadows varies according to moisture content, elevation, and soil structure (Clary 1999; Keeley et al. 2003). At the level of the individual grazin g impacts are a function of timing and intensity. Plants are particularly more su sceptible to defoliation during the boot stage of their development (Clary 1995), that is th e portion of the repr oductive cycle at the onset of flowering right before the seedhead emergences from the sheath of the flag leaf. Grasses and forbs produce fewer inflorescences when defoliated just before or during flowering (Edwards 1985; Olson-Rutz et al. 1996). Many montane grasses and forbs also tend to show decreased stem length (Mueggler 1967 and 1972: Trlica et al. 1977; Stout et al 1980; Stout and Brooke 1987; Olson and Rich ards 1988) and leaf growth in the next 31

PAGE 40

growing season after grazing (Mueggler 1972 and 1975; Edwards 1985). Grasses grazed during flowering produce fewer tillers per unit area (Storer et al. 1980, Stout et a l. 1981) and have lower tiller replacement rate (Ols on and Richards 1998). Heavy or prolonged grazing, even once, if at an inopportune time, can result in effects that carry over for a year or more (Clary 1995; Olson-Rutz et al. 1996) 1.5.1 Plant species composition an d Exotic species invasions Given these generalizations, it is important to remember that measures of plant responses, such as decreased tiller growth a nd stem length, indicate widely varying levels of damage depending of the species being evaluated (Olson-Rutz et al 1996). Most montane meadow Carex species are considered grazing resistant due to their rhizomatous growth habit. Therefore, a large amount of defoliation of these species may not be very significant on a population scale. Other sp ecies may be heavily impacted by even a single grazing event (Clary 1995). The presence of grazers may favor the establishment and growth of some species and decrease the presence of others. Ther efore, a useful con cept for evaluating the condition of rangelands is the proportion of sp ecies designated as increasers, decreasers and invasives present at climax state. In a climax condition meadow, the abundance decreaser species tends to dec line with overgrazing. Similarl y, increasers species exhibit greater abundance in response to overgrazi ng, at least initially. The proportion of increasers in relation to decreasers in a given meadow ma y indicate meadow degradation. Climax vegetation in pristine meadows genera lly has a high proportion of decreasers and relatively few increasers. Degraded mead ows have the opposite compositions; many 32

PAGE 41

increasers and few decreasers (Dyksterus 1949 ). Ratliff (1985) has compiled a list of montane meadow increasers, decreasers and inva sives specific to Sierran meadows. An abridged version of this list can be seen in appendix A and includes only those species observed in this study. The classification of these species can vary under certain conditions. Some species repr esent a greater deviation fr om climax vegetation than others. Therefore, the presence or abundance of either increasers or decreasers does not necessarily indicate overgrazing and use of this system can be used only as a general guide (Ratliff 1985). 1.5.2 Defoliation Measures of defoliation impacts on meadow plant growth is often determined through experiments in which meadow foliage is clipped to uniform heights (Clary 1995; Martin and Chambers 2001; Klus e and Allen-Diez 2005; Berlow et al. 2002). Meadows tend to lend themselves to clipping analysis as cattle tend to graze vegetation to uniform heights in montane meadows (Bartolome 1984). The exact impact of different grazing in tensities on a given meadow appears to depend heavily on the height of the water table and the ti ming of grazing. A study of simulated grazing in several meadow co mmunity types found different responses between Kentucky bluegrass (Poa pratensis L.) dominated communities and Sedge ( Carex spp.) dominated communities (Clary 1995). Poa pratensis, a non-native graminoid, is more prevalent in dry meadow types while the common dominant sedge species Carex nebrascensis characterizes wet meadow types (Martin and Chambers 2001). In grazed meadows, Kentucky bluegr ass communities maintained or increased 33

PAGE 42

above groun d biomass while sedge dominated communities maintained or lost biomass. In sedge dominated communities, heavy grazing during growing season resulted in significant biomass loss in subsequent years. Heavy grazing in this context refers to grazing once annually to a 5 cm stubble heig ht, 10 cm in late summer, or a general utilization rate of 30% or greater (Clary 1995). Martin and Chambers (2001) proposed that P. pratensis increases with grazing because it can rapidly increase its growth rate to take advantages of disturbances that remove neighbors and increase space. In the southern Sierras, the sedge species, C. aguasta, dominates the sedge dominated meadow communities that are dominated by C. nebrascensis in the northern reaches. C. aguasta, however, appears to fill a si milar ecological role as C. nebrascensis. Other meadow species appear to be similar in both areas (Allen-Diez 1991). Poa pratensis and Deschamsia cespitota, a native perennial bunchgrass, dominate similar meadow sites and are consid ered valuable range species. P. pratensis is considered an increaser species while D. cespitota is considered a decreaser. Among mesic meadows D. cespitota tends to dominate meadows on the higher end of the soil moisture gradient. P. pratensis has recently increased its ra nge, however, it is unclear what role grazing plays in this increase. While grazing significantly decreased the biomass potential of both species, it did not se em to give the competitive advantage to Poa over D. cespitota However, larger scale changes to meadow hydrology, such as lowering of the water table due to cattle in fluenced stream incision would favor the establishment of P. pratensis over D. cespitota. Simulated grazing of P. pratensis appeared to release Deschampsia from competitive inhibition. Therefore limited grazing may actually aid in th e establishment of D. cespitota (Kluse and Allen Diez 2005). 34

PAGE 43

1.5.3 Preferential Grazing Preferential grazing refers to the preferen tial selection of certain forage species by herbivores. At low levels of herbivory, such as those exerted in meadows by moderate abundances of native grazers, th e changing palatability of pl ants over the course of the season effectively prevents pr eferential grazing (Ratliff 1985). Grazing, if continued over a prolonged time period or frequently repeated, can impact the population of a preferred species e nough to severely increase its scarcity. Therefore, it is important for managers to consider the grazing habits of different animals when outlining or designating allotments. Cat tle prefer to graze grasses, browses, and forbs, in that order. Sheep also fully utili ze grasses but also utilize forbs more so than cattle. Among native grazers, big horned sheep prefer grasses and grass-like plants but also utilize forbs in the spring and summer. Deer, which have a very small dietary overlap with cattle, prefer forbs and brow ses year round (Ratliff 1985). Moisture and topography also tend to dictate grazing preferences. Cattle tend to congregate in the dryer portions of meadows early in the s eason but will move to the wetter meadow portions later in the season as meadow aridity increas es (Pattee 1973). Sheep, on the other hand, tend to avoid wet meadow portions (Ratliff 1985). Packstock (i.e. mules and horses) preferentially graze graminoids over forbs. Preference for forbs is a function of hunger and it appears that forbs are only grazed after gaminoids have been sufficiently depleted (Strand 1979, Olson-Rutz et al. 1993) Mules show a greater degree of species prefer ence than horses and may make larger contributions to local species depletion (Ratliff 1985). 35

PAGE 44

1.5.4 Trampling Tram pling can cause local changes to soil structure, content and dynamics which may favor species adapted to those conditi ons (see section 1.3). Very wet sphagnum dominated meadows are especially vulnerable to cattle trampling. Cattle hooves tend to break up the sod leading to ev entual vegetation death and the formation of mud holes (Ratliff 1985). High soil compaction impedes the growth and development of the fibrous roots of perennial grasses. In these conditions plants with strong taproots or shallow rooted annuals may become dominant. Changes to pH, due to consistent trampling of human and livestock trails may work to favor species that are especially adapted to acidic soils (Leonard et al. 1968; Lockaby 1984). However tr ampling also drives certain beneficial processes in meadows. People a nd animals can transport seeds and rhizomes on their boots and hooves, contributing to ge netic flow and effectively reintroducing excluded plants. Also some compaction can prevent frost damage (Ratliff 1985). 1.5.5 Effects on Biodiversity and Invasibility Moderate grazing generally creates conditi ons for highest species diversity (Olff and Richie 2003) On the other hand, overg razing tends to simplify meadow communities and lead to the establishment of stands of undesirable fora ge species (Dorrough et al. 2007). Montane rangelands in the Sierras have not been as heavily impacted by invasive species as rangelands at lower elevations. Hi gh species richness did not appear to affect the invasibility of these ecosystems. Both species richness and richness of invasive species tend to decrease as elevation increa ses, suggesting that invasive species are 36

PAGE 45

lim ited by the same factors as native species (Keeley et al. 2003) Analysis of grazed land and land that has been protected from grazing for nearly 100 years show only sight increases in invasive species presence. This suggests, invasive plants, once established, exhibit some degree of persis tence, even after grazing pr essures have been removed (Keeley et al. 2003). Lyons and Swartz (2001) proposed that this persistence of invasive plants may be due to a self reinforcing system of alternative stabi lity. They propose that communities high in non-native species themse lves resist re-invasion by native species. However, among meadows located at simila r elevations, it is unclear how grazing influences biodiversity. Another study of long term grazing protection found that, after 65 years, plant diversity inside and outside of cattle proof enclosures di d not differ among 11 of 16 enclosures in the Great Basin (Curtoise 2004). Sierran meadows do not experience the same connectivity or climatic pressures as arid, lowland ranges like those found in the Great Basin. Meadows are ge nerally isolated from each other by abruptly heterogeneous vegetation. It is possible that meadows bi odiversity may react differently to grazing exclusion as it would be more difficult for depl eted seed banks of preferentially grazed species to be replenished. It is clear that grazing tends to decr ease the cover and ri chness of woody species including willows ( salix sp.). Woody species are impor tant habitats for wildlife. Meadows with a better develope d shrub layer support a grat er diversity of birds and mammals (Woudenburg 1999). Meadows containi ng willows are important habitats for the rare willow fly catcher (see section 1.4.1). 37

PAGE 46

1.5.6 Shrub and Tree Expansion In the last 1 00 years, since the inception of grazing controls in the Sierra Nevada, expansion of woody species into meadows ha s been a widespread agent of meadow change and disappearance. Yosemite and Sequoia National Parks lie to the North and South of my study site, respectively. Analysis of historical photos show that meadows, in both parks, have lost area to tamarack pine invasion (Pinus contorta var. murrayana ) (Vankat and Major 1978; Vale 1987). In Yosemite National Park, pine invasion appeared to be a problem associated with dry meadows and the upland borders of large meadows. Wet meadow lands often, but not always, were free of trees. Grazing is thought to contribute to tr ee invasion in these meadow s (Vale 1977; Dunwiddle 1977; Vale 1981; Vale 1987) Grazing may drive these changes thr ough a variety of mechanisms. Heavy grazing can reducing plant cover and crea te mineral soil surfaces which promote tamarack seed germination (Vale 1987). Grazing reduces combustib le forage thereby decreasing frequency of burni ng and increasing the survival of woody species (Leopold 1924; Milchunas and Lauenroth 1993; Vavra et al. 1994; Van Auheb and Bush 1997). Livestock presence can contribute to eros ion and change meadow hydrology. Erosion driven incising of stream banks, soil comp action, and other processes that increase meadow aridity produce favorable conditions for woody species encroachment (see chapter 1.2; Platts 1979; Magil ligan and McDowell 1997; Kirchner et al. 1998). Livestock may aid in dispersing seeds into the meadows through their dung and on their bodies (Browne and Archer 1988; Browne and Archer 1999). Polly (1997) posits that 38

PAGE 47

elevated CO2 levels created by high density of livestock may increase shrub establishment as well. Meadow moisture determines how cattle grazing impact woody species encroachment. Sagebrush ( Artemisea sp.) tend to encroach on arid meadows and those meadows with large dry portions. These meadow s are most characteristic of the southern and eastern Sierra. While encroachment is most prevalent in arid meadows and meadow portions, Barlow and colleges (2002) found in a series of clipping experiments that the establishment of Artemisia rothrokii Grey is highest in mesic/moist meadow sites given sufficient levels of disturbance. Undi sturbed mesic sites effectively prevented germination and inhibited seedling survival. At very wet sites grazing is key to woodyshrub mortality and prevention of succession into the woody-sh rub state. At very dry sites, simulated herbivory neither inhibited or encouraged woody shrub establishment. However, livestock driven changes to mea dow hydrology which lead to decreased soil moisture content may shift vegetation interactions in once-wet sites in favor of the shrub establishment (Barlow et al. 2002). Pocket gophers mounds have been implicated in creating suitab le vegetation gaps in which tamarak pine seedlings may become established (Buchanan 1972). Ratliff (1987) suggests that overgrazing by domestic cattle can aggravate the effects of pocket gophers in montane meadows. Also, overgra zing may expose seed predating rodents to greater predation themselves thereby reducing the amount of tamarack pine seeds consumed by these rodents (Ratliff 1987). Irrespective of the impact of grazing disturbance on the continuation of shrub impact, some initial disturbance must create instability in the mead ow system for shrub 39

PAGE 48

40 expansion to commence. Grazing may trigger in stability by increasing the survival rates of a few seedlings that become established in the meadow. However the actual driver of shrub expansion is ambiguous, geological and climatic factor s may contribute as well. Vale (1987) observed that shr ub expansion in the meadows of Yosemite National Park did not begin until after the cessation of sh eep grazing. Therefore seedlings could not survive trampling but did favor the conditi ons created after grazing control in 1905. Muller and Woolfenden (2003) poi nt out that recorded cases of shrub expansion also coincided with the climatic changes due to the end of t he little ice age in 1900. Therefore shrub encroachment may have been triggered by climate change that was unrelated to grazing regulations.

PAGE 49

2. Materials and Methods Data for this study were collected at 10 different mid elevation montane meadows in the area surrounding Shaver Lake, Calif ornia (Fig 2.1). Four meadows were designated grazed and the other six, ungrazed. Very few sites in the Sierra Nevadas remain entirely ungrazed, even in very remote reaches (McKelvey and Johnston 1992). It may be more appropriate to refer to ungrazed study sites as grazing protected. All meadows in this study were grazed at one point in time. The only possible exception is Stevenson meadow. Stevenson meadow is a restored meadow that was constructed by the Southern California Edis on Company in 1987 over the site of a blast debris pile. Because this meadow did not act ually exist before the removal of grazing, it is arguable that this meadow had never b een grazed. However this meadow is also disturbed. Several extensive topography ch anges have been made to correct the hydrology of Stevenson meadow since it s construction (Tansky 2008 personal communication). Overall this meadow is much dryer and with much sparser vegetation in comparison with the other meadows surveyed. Sites were considered ungrazed/grazing protected if a concerted effort to prevent cattle grazing had been in place more than 20 years. Most ungrazed meadows were located on private property belonging to S outhern California Edison. Ungrazed land was fenced to exclude cattle and domesticated liv estock in the 1980s as per an agreement with the U.S. Forest Service, from which th e land is leased. Fences around the land were maintained until the last few years, at which time some were damaged or fell into disrepair (Bird 2008 personal communication). 41

PAGE 50

Shaver Meadow has a pack animal rental station situated upo n the Northern end. While pack animals are routinely moved thr ough the meadow on designated trails, they are not allowed to graze (Bir d 2008 personal communication). Th erefore, this meadow is designated ungrazed. The Dinky Meadows area, on the south side of Shaver Lake, contains many meadows that fall under USDA jurisdiction and are grazed. Several meadows close to the lake are still under the stewardship of Southern California Edison, and are ungrazed. Meadows were considered ungrazed if the three following conditions were met: All cattleproof fences were intact and the meadow was in an area designated for cattle exclusion, The meadow is geographically isolat ed from grazed meadows (inaccessible to livestock because of surrounding features such as bodies of wa ter, high ridges or peaks, or densely forested areas of land, and, No traces of cattle grazing are visi ble (dung, hoof prints or highly trampled ground). Grazed control meadows included all those on U.S. Forest Service land and those on Edison land that showed obvious signs of ca ttle invasion. Meadows we re located between 1550 and 1850 m in elevation. Vegetation surveys were conducted in 10-25 quadrats per meadow depending upon meadow size. In each meadow, at least 10 quadrats were measured regardless of meadow size. In meadows much larger th an 10 ha, another quadrate was added for approximately every hectare over the 10 ha base line. Quadrat locations were chosen by 42

PAGE 51

Mushroom Rck I nset Shaver l ake I nset Figure 2. 1 Relief map of study Area, a red bulls eye indicate meadow center point locations (base map courtesy of U.S.D.A. forest service, Modified using ArcGIS software. 43

PAGE 52

obtaining numbers from a random number generator and moving the corresponding amount of meters down the north-south or eas t-west axis of the meadow. Each quadrat was one square meter in area. Non vas cular plants, fungus, and lichens were not included in the survey. GPS coordinates were recorded at the site of each quadrat. Plant identifications were conducted both in the field and the laboratory. Those plants that had distinct forms and inflores cences and could be easily identified were identified in the field. Those that could not be identified in a tim ely manner were given a letter label and a sample was collected for later species determinati on. Flowering herbs and the few woody species that were encountered were identified using the Sierra Nevada Field Guide by John Muir Laws (2007). Most grass species and some herb species not included in the Laws field guide were iden tified by dichotomous key using the Jepson Manual of Higher Plants of California (1993). Species lists from Southern California Edison were also used to gui de species determination. The distances used in these analyses were determined using the Geographical Imaging Software, ARCGIS, and its supporting programs. GIS refers to a number of programs in which physical information is coupled with geographic designations for the purpose of visualization, manipula tion, and analysis. Information is stored in the form of layer files. A layer consists of related geographic information in the form of raster files or shape files and explanatory designations called attributes. For example, a layer may describe public roads in a given city. Each road may have a number of attributes including name, number of lanes, maintenance status or ownership. These layers may be combined to create maps and to relate different sets of spatial information. 44

PAGE 53

In this study, the density of roads represents both direct disturbance as well as an approximation for other anthropoge nic disturbance. While roads impart a certain amount of disturbance into the mead ow by their physical presence they also represent the potential for a myriad of othe r disturbance factors, i.e. ge neral commercial areas, logging, construction, and hydropower development. A meadow with nearby roads allows greater access to recreational hikers. Finally roads introdu ce vectors for exotic plant propagules as seeds can be introduced in tire treads and in cargo. A shape file was created containing coordinates for all the meadow quadrats in the study (Fig 2.2). A road layer was constructe d by digitizing a paper map of the Shaver area into arc GIS. Road s were classed by intensity of development on an ordinal scale from 1 (secondary unpaved roads) to 4 (highw ays). Road classes and their criteria are detailed in table 2.1. Buffer areas, with 1000m radii were drawn around the data points representing meadow quadrants. Buffers were dissolved in such a way that any overlapping buffers were combined into single sh ape file (fig 2.2). The length of all roads within these buffer areas was calculated. To account for the length of road in a 1000 m proximity of the meadow quadrat as well as the potential impact of a more heavily developed and well-traveled road, a road index was form ulated (Equation 2.1). ir=( lc)/ a Equation 2. 1 road index where l= le ngth of a road in the buffer area c=class value for that road and a=area of the buffer Similarly, a trail index was formulated to accoun t for the proportion of trails in the buffer area (Equation 2.2). 45

PAGE 54

it=( l)/a Equation 2. 2 road index where l= le ngth of a trail in the buffer area, and a=area of the buffer. Road classes do not apply However, trails were not classed and the c lass term is excluded from the equation. Some designated hiking trails run along roads; therefore, only stand-alone trails were measured. Meadow distances from campgrounds we re measured from the center point of the meadows to the center point of the meadows to the center of each campground. Table 2. 1 Criteria for road classification Road class Class Value Criteria highways 4 Paved two lane roads, open year round to the public surfaced 3 paved, one or two lanes, residential or private primary seasonal 2 one lane unpaved roads, generally gated to the pu blic but used heavily by workers, well maintained and graded often secondary seasonal 1 one lane unpaved, gated to the public periodically maintained Biodiversity was measured using the Sh annon-Weiner index (Equation 2.3). The Shannon-Weiner biodiversity i ndex takes into account both species richness and the species evenness, the equali ty of species representation in a study area (Shannon and Weaver 1959). s H = (Pi ln Pi) i=1 Equation 2. 3 Computing formu la for the Shannon index where ni =The number of individuals in species i or the abundance of species i S =The number of species (aso called species richness), N =The total number of all individuals, pi =The relative abundance of ea ch species, calculated as the proportion of individuals of a given species to the total number of indi viduals in the community: ni/N (Shannon and Weaver 1959) 46

PAGE 55

Species abu ndance was measured by taking the number of species in each quadrat, dividing 1m2 by that number resulting in the term aq where q is the quadrat number. Each species was considered as having aq as the amount of ground cover for that quadrat. Then aq was summed over all quadrats in the mea dow. Biodiversity and Exotic cover in relation to meadow grazing status was analyzed using a parametric two-tailed t-test. Exotic cover in relation to roads, trails, and campground proximity were analyzed by linear regression. 47

PAGE 56

Fig 2. 2 Map insets from figure 3.2 depicting the Shaver Lake and Mushroom Rock study areas. Road density was measured inside light teal bu ffer areas; Distance to campgrounds was measured between red meadow centers and blue campground ic ons using special analysis functions in the arc GIS software package (ArcGIS 2009) 48

PAGE 57

3. Results Lists of the m eadow plant sp ecies and statistical tables are listed in appendix A. 3.1 Biodiversity Mushroom Rock Meadow (Fig 3.1) was unus ual in relation to the other meadows because it was situated at a much higher el evation than the other meadows and it was unusually large. This meadow also exhibi ted exceptionally high species richness and biodiversity. I decide to exclude mushroom meadow as it was not representative of the sample (fig 3.3). Figure 3. 1 Mushroom meadow (photo by Joseph Dronchi, July 2008) 49

PAGE 58

Mean Shann on biodiversity in grazed m eadows was slightly lower in ungrazed (mean Shannon index=3.028) than grazed meadows (mean Shannon index=3.293). A two tailed t-test of mean biodiversity showed a significant di fference between grazed and ungrazed meadows (t=2.04, df=5.65, p<0.09). U ngrazed meadows exhibited a larger variability in biodiversity values. Biodivers ity values in grazed meadows, on the other hand, occupied only a narrow range of high values (fig3.2). Biodiversity, among the 10 meadows was not significantly correla ted to the trail index (r=0.0005, p 0.989) or road index (r=-0.061, p 0.867). Nor was there a significant correlation of biodiversity to the proximity of meadows to campgrounds (r=0.319, p 0.369) (figure 3.4 a, b, and c). maximum value 75th percentile mean median 25th percentile minimum value Figure 3. 2 Box plot of Shannon Biodiversity values of grazed and ungrazed meadows, Mushroom Meadow is excluded 50

PAGE 59

Figure 3. 3 Box plot of Shannon index values of grazed and ungrazed meadows, Mushroom Meadow is included for purpose of comparison 51

PAGE 60

A. 2.5 2.7 2.9 3.1 3.3 3.5 3.7 0100020003000400050006000 proximity of campgrounds (m)shannon biodiversity B.2.5 2.7 2.9 3.1 3.3 3.5 3.7 01234 trail index (1/m)shannon biodiversity C.2 2.2 2.4 2.6 2.8 3 3.2 3.4 3.6 3.8 49141924road indexshannon biodiversity index Figure 3. 4 scatter plots of shannon biodiversity in realtion to A) proximity to campgrounds in meters B) Trail index (trail index increases with increasing trail density) and C) Road index (road index increases with increasing road density and road development) 52

PAGE 61

3.2 Exotic cover The presence of exotic species in the meadows, as a whole, was minimal. No more that two different speci es were found in any given m eadow. Exotic species cover was also minimal. No meadow exhibited an ex otic cover greater than 8%. No significant correlations where observed in relation to grazing, proximity to campgrounds, road index, or trail index (Figures 3.5 and 3.6). Figure 3. 5 Exotic plant species ground cover in grazed and ungrazed meadows 53

PAGE 62

A.0% 1% 2% 3% 4% 5% 6% 7% 8% 9% 0100020003000400050006000 proximity to campground in meterspercent exotic cover B.0% 1% 2% 3% 4% 5% 6% 7% 8% 9% 0510152025 road indexpercent exotc cover C.0% 1% 2% 3% 4% 5% 6% 7% 8% 9% 0123 trail index percent exotic cover 4 Figure 3. 6 Scatterplots of exotic plant cover in rel ation to A) proximity to campgrounds in meters B) road index and C) trail index 54

PAGE 63

4. Conclusions and Discussion 4.1 Discussion 4.1.1 Biodiversity Meadow which were grazed and those meadows that were grazing protected significantly differed in biodive rsity but not exotic cover. It might be concluded, therefore that the presence of livestock, at least at the stocking rate that is currently enforced, increased biodiversity of the ten m eadows surveyed in this study. However, if one where to compare meadows with higher stocking rates a signi ficant decreased in biodiversity might be fou nd (Mueggler 1967). Biodivers ity variability in ungrazed meadows is partially explained by a larger sample size of ungrazed than grazed plots. Larger sample samples are more likely to contain a wider range of biodiversity values. Cattle may change biodiversity by three mechanisms; selective grazing (Ratliff 1985, Beck and Peek 2005), changes in hydrology (Allen-Diez 1991, Vale 1981) and facilitation of exotic plant establishment (Harrison 1999). Intermediate levels of disturbance result in the highest biodiversity (Grime 1973). Some grazing can open vegitation gaps and create resource islands thereby allowing certain plants to be established that would, othe rwise, be out competed. At very high stocking rates, grazing is expected to lower biodiversity as some plants are selectively grazed out and freque nt disturbance insure s only stress tolerant plants can flourish. Cattle provide intermedia te levels of disturbance in grazed meadows to maintain consistently high biodiversit y, regardless of area or hydrology. Ungrazed meadows may or may not exhib it high biodiversity without catt le as an added disturbance factor. 55

PAGE 64

Evidence of grazing-driven changes in biodiversity m ight be detectable by monitoring meadow hydrology and the proporti on of increasers and decreasers over time (Platts 1979; Ratliff 1985). However, this kind of analysis is not possible with the snap shot data collected for this project. Even if the sample size of the meadows was increased, such an analysis would de difficu lt without long term monitoring. This is because pre-grazing vegetation composition va ries widely according to hydrology and origin. A dry ungrazed meadow may have mo re increaser species than a grazed wet meadow (Stohlgren 2002). Therefore, to make a snap shot analysis of increasers and decreasers in grazed and ungrazed meadows, one would need to find two meadows that closely resembled each other in proportions of hydrologically similar meadow sites. The proximity of roads and campgro unds are indicators of non-livestock disturbance. Campgrounds repr esent points from which recrea tional hikers may travel to and around meadows (Boyle and Sampson 1985). In my study area recreationists also tended to exacerbate cattle intrusion by dismantling fences intended to exclude cattle from sensitive areas and grazing excluded m eadows (personal observation, Bird personal communication). The proximity of campgrounds was not significantly related to biodiversity or exotic cover. Trail index was also unrelated to biodiversity. These data suggests that non-consumptive, recreatio nal camping and hiking, alone, does not significantly impacts meadow vegetation composition. These activities are generally considered acceptable and low-impact uses of public land. However, other kinds of disturbance such as commercial developm ent and road construction can accompany the establishment of tourism and recreation sites (Rudzitis and Johnson 2000). 56

PAGE 65

W hile I attempted to ensure the meadows included in this study were similar in most respects (i.e. elevation, size, and mois ture) it was not possible to control for all factors. Some of the larger meadows contai ned sites that differed widely in moisture levels, with boggy areas waist deep with stan ding water, in the center, and very dry sparse edges. It would be expected that m eadows with a wider range of microhabitats are expected to have a higher biodive rsity simply by virtue of the nu mber of niches available. If meadow size were to have been included in this study one might expect a correlation between large meadows that contain many di fferent microhabitats and high biodiversity. Conservation priorities should be established to reflect the contributions of large meadows with large swaths of diverse microhabitats to th e Sierran ecosystem. Large meadows tend to have high species diversit ies and are vital to willow fly catcher populations (. Therefore, the protection of a few large meadows may be of greater benefit to wildlife populati ons that many of the small and isolated meadows. 4.1.2 Exotic Cover The presence of exotic plants in all of the meadows was minimal. Exotic invasion is not as considerable a problem in the mont ane elevation as in the lower oak woodlands. While exotic cover never exceeded 9% in my study plots, exotic species comprise the dominant vegetation in nearby lowland oak savannas (Safford and Harrison 2001; Keely et al. 2003). The most common exotics observed were Poa pratensis and Taraxacum officianale Wigg Poa pratensis is of Eurasian origin; howev er, its introduction predates European exploration. It is unknown how long introduction of Poa pratensis predates western expansion. Exotic cover was so low, and invasives so infrequently encountered, 57

PAGE 66

in all m eadows, that a detection of 1% exotic ground cover versus 8% exotic ground cover may not indicate a substantial difference in exotic plant success. All areas of the Sierras have been grazed at some point and exotic introductions have at least some degree of persistence in Sierran ecosystems (Safford and Harrison 2001). It may be concluded that exotic introductions can be impacted very little by curr ent land uses. None of the meadows investigated in this study were located in pristine or even particularly remote regions, therefore, one may assume that su fficient propagule pressure exists for the introduction of exotics into these meadows. The question arises as to why some meadows showed some exotic presence and ot hers showed none. The key to exotic persistence has to lie in the existence of fa vorable habitat for invasive establishment and growth. Favorable conditions may exist in m eadows with large swat hs of mesic meadow sites as opposed to those that are mostly very wet (Dwire et al. 2004). 4.2 Conclusions Montane meadows in the Sierra Nevadas are vital habitats for plants and wildlife. The maintenance of diversity and vital eco system processes in these habitats are important to the overall health of the Sierra Nevadas. The conservation of these habitats should be a priority for conservation mana gers because of the great abundance and diversity of endemic species in the California Floristic Province. Despite the delicate and important nature of these habitats meadows must weather the stress of multiple uses. Meadows contain valuable fora ge for livestock as well as wildlife. They act as vital watersheds and are valued by outdoors enthusiasts. While a great amount of research has addressed the impact of grazing on the ecosystems of the west it is difficult to extrapolate 58

PAGE 67

these r esults to specific habitats. Grazing has different effects given a myriad of factors; soil moisture, meadow topography, prior use, time of year, and stocking density of livestock. While plant biodiversity is important to the survival of endemic species and the maintenance of ecosystem stability (Tilman 1996), it is by no mean s the only factor in management decisions concerning montane m eadow use. Even slight changes to vegetation structures dramatically influence obligate meadow wildlif e populations. Some species benefit from habitat changes induced by grazing while others are harmed. Therefore, specific management goals s hould be established before management decisions are made. It is important that managers establish whether the meadow in question contains sensitive wildlife species such as willow fly catcher or pika. Many meadows do not harbor components necessary for these sensitive species and therefore might be moderately grazed without detriment to sensitive wildlife. Meadows that are already geologically unstable ma y experience greater impacts from grazing as the added disturbance may completely destabilize the m eadow and initiate successional processes. Data obtained in this study is far from complete. The impact of grazing on plant populations, beneficial or unfavor able, is influenced by a number of factors. To better understand the impact of livestock distributions on plant populations in Sierran meadows, managers should focus on assessing meadow ch ange over the long as well as differences in grazed and ungrazed meadows. Conditi ons within individual meadows can very widely: larger meadows tend to have higher biodiversity due to increases in microhabitats (Dwire et al. 2004), wet meadows have lower diversity than mesic meadows and dry meadows (Dwire et al. 2006). These differences make comparisons among individual 59

PAGE 68

60 meadows difficult. Long-term monitoring of a few meadows may provide greater insight into meadow dynamics than comparisons of grazing in many different meadow types. The results of this study suggest that cattle grazing, at th e current stocking density, does not adversely affect biodivers ity or exotic species presence, and even slightly increases it in grazed meadows. Interm ediate levels of distur bance, such as those created by low densities of cattle, potentially increase species richness. However, ungrazed meadows may exhibit high biodive rsity regardless of cattle absence. The meadows of the Sierra Nevadas are sma ll, albeit vital, pieces of a diverse and incredible ecosystem. While these little meadows make up such minute portions of the landscape, their contribution ecosystem processes in the Sierras is great. As more becomes understood about the dynamics of these habitats, land manage rs may be better equipped to balance the needs of the plant, wildlife, and human populations. It is important that we maintain this balance to ensure the preservation of the Sierra Nevadas long after the transient demands of our generations have passed.

PAGE 69

Appendix A: Species List All plant species observed in the course of my survey are liste d below; these are only those species that fell within the samp le quadrat. Some species were observed coincidental ly but not counted, and as such, are not listed. Origin Grouping N= native I=increaer E=Exotic D= Decreaser Latin Name author Common Name Origin g Grouping Achillea millefoliumYarrowN Achnatherum occidentalisWestern Needle grassN Agastache urticifoliaNettle Leafed Giant HyssipN Agroseris glauca N Agrostis albaRedtopN Agrostis idahoensisIdaho BentgrassN Angelica breweriBrewer's AngelicaN Aster occidentalisWestern Meadow AsterNI Brodiaea elegans ssp. elegansHarvest BrodiaN Camassia quamashSmall CamasN Carex anthrostachys ND Carex integrasmooth beaked sedgeND Carex jonsii ND Carex nebrascensisNebraska sedgeND Carex vesicariaBlister SedgeN Carex vesicaria var. majorWestern Inflated SedgeN Castilleja mimiata N 61

PAGE 70

Ceanothus diversifolius Pine Mat Ceanothus N Cuscuta californica Hook and Arn. California Dodder N Danthonia californica Bol. California Oatgrass N I Deschampsia cespitosa ssp. cespitosa (L.) P. Beauv. Tufted Hairgass N D Dodecatheon jeffreyi Van Houtte Jeffery Shooting Star N I Elymus glaucus Buckley Blue Wildrye N I Elymus multisetus M.E. Jones Big Squirril tail N Elytrigia intermedia(Host) Barkworth & D.R. DeweyIntermeadiate Wheatgrass N Equisetum arvense L. Field Horsetail N I Fragaria vesca L. Wild strawberry N Fragaria virginiana Duchesne Mountain Strawberry N Gayophytum eriospermum Coville Coville's Groundsmoke N Geranium richardsonii Fisch. & Trautv. Richardson's Geranium N Geum macrophyllum Willd. Largeleaf Avens N Glyceria elata (Lam.) Hitchc. Fowl Mannagrass N D Helenium bigelovii A. Gray Bigalow's Sneezeweed N I Heracleum lanatum Bartram Cow Parsnip N Hordeum brachyantherum Nevski Meadow Barley N I Hordeum jubatum L. Foxtail barley N Hypercom perforatum L. Klamath weed E Ivesia lycopodioides A. Gray Clubmoss Ivesia N I Ivesia sericolonea Clubmoss Ivesia N Juncus effusus L. Common Rush N Juncus nevadensis S. Watson Sierra Rush N D 62

PAGE 71

Kackiella braviflora Lindley Pestemon N Lathyus jepsonii E. Greene Jepson's Pea N Linum lewisii Pursh Lewis's flax N Lotus oblongifolius (Benth.) E. Greene Meadow Lotus N I Lotus purshianus(Benth.) Clements & E.G. Clements Spanish clover N Lupinus albifrons Benth. Bush Lupine N Lupinus covillei covillei E. Greene Shaggy Lupine N Mimulus guttatus D.C. Seep-spring Monkey Flower N Mimulus primuloides Benth. Primrose Monkey Flower N I Mimulus tilingii Regel Mountain Monkey Flower N I Monardella lanceolata A. Gray Mustang Monardella N Pedicularis attollens A. Gray Little Elephant Head N I Penstemon procerus var. formosus (Nelson) Cronq. Little Flowered Pestemon N D Penstemon rydbergii var. oreocharis (Greene) N.H. Holmgren Herbatous Pestemon N Perideridia parishii (J.M. Coult. & Rose) A. Nelson & J.F. Macbr. Parisher's Yampa N D Phleum pratense L. Timothy N D Poa pratensis L. Kentucky Bluegrass E I Polygonum bistortoides Pursh American Bistort N I Potentilla drummondii (S. Watson) Ertter Drummond's Cinquefoil N Potentilla flabellifolia Hook. ex Torr. & A. Gray Fan Leaf Cinquefoil N I Potentilla glandulosa Lindl. Sticky Cinquefoil N I Potentilla gracilis v. fastiata Douglas ex Hook. Slender Cinquefoil N I Potentilla grayi S. Watson Gray's Cinquefoil N Raillardella scaposa A. Gray Raillardella N Ranunculus occidentalis Nutt. Buttercup N 63

PAGE 72

Rhus trilobata Nutt. Skunkbrush Sumac N Rudbeckia californica A. Gray California Coneflower N Rumex crispus L. Curley Dock N Salix exigua Nutt. Narrow Leaf Willow N D Scirpus microcarpus J. Presl & C. Presl Panicled Bull Rush N I Sidalcea glaucescens Greene Checker Bloom N Sisyrinchium idahoense E.P. Bicknell Idaho Blue Eyed Grass N Solidago canadensis L. Canada Goldenrod N I Solidago spp. goldenrod unknown species N Stachys albens A. Gray Whitestem Hedgenettle N Taraxacum officinale F.H. Wigg Common Dandelion Exotic Thysanocarpus curvipes Hook. Sand Fringepod N Trifolium longipes Nutt. Longstem Clover N I Trifolium pratense L. Red Clover N Trifolium willdennovii Spreng. Tomcat Clover N Triteleia ixioides ssp. anilina (Greene) Lenz Prettyface N Triteleia ixioides ssp.scabra (Greene) Lenz Prettyface N uk I uknown uk E uknown uk 12 uknown uk 13 uknown uk 16 uknown uk 17 uknown uk 18 uknown uk 19 uknown 64

PAGE 73

uk 19 uknown uk 21 uknown uk 4 uknown uk 8 uknown uk B uknown uk Q uknown uk T uknown uk V uknown uk W uknown Veratrum californicumDurandFalse HelleboreNI Viola macloskeyiLloydMacloskey's violetNI 65

PAGE 74

References Ada ms SN. 1975. Sheep and cattle grazing in forests: A review. J Appl Ecol :143-52. Airola DA. 1986. Brown-headed cowbird parasitis m and habitat disturbance in the sierra nevada. The Journal of Wildlife Management :571-5. Allen B. H. 1989. Ten years of change in si erran stringer meadows: An evaluation of range condition models. Proceedings of th e california riparian systems conference: Protection, management, and restoration for the 1990's, edited by DL abell. 102 p. Allen BH. 1987. Forest and meadow ecosystem s in california. Rangelands 9(3):125-8. Allen-Diaz B, Berkeley U, Barrett R, Hunt singer L. 1999. Sierra Nevada ecosystems in the presence of livestock. A Report to the Pacific Southwest Station and Region.USDAFS Allen-Diaz BH. 1991. Water tabl e and plant species relations hips in sierra nevada meadows. Am Midl Nat 126(1):30-43. Anderson MK and Moratto MJ. Native Amer ican land-use practices and ecological impacts. Report NR 2. 187-206 p. Bai Y, Don Thompson, Broersma K. 2000. Ea rly establishment of douglas-fir and ponderosa pine in grassland see dbeds. J Range Manage 53(5):511-7. Bailiff RD. A meadow site classification for the Sie rra Nevada, california. Bartolome JW. 1987. California annual grassland and oak savannah. Rangelands :122-5. Beck JL and Peek JM. 2005. Diet composition, forage selection, and potential for forage competition among elk, deer, and livestock on aspen-sagebrush summer range. Rangeland Ecology & Management :135-47. Beever EA, Brussard PF, Berger J. 2003. Patte rns of apparent extirpation among isolated populations of pikas (ochotona princeps ) in the great basi n. J Mammal :37-54. Berlow EL, D'Antonio CM, Reynolds SA. 2002. Shrub expansion in montane meadows: The interaction of local-scale disturbance and site aridity. Ec ol Appl 12(4):1103-18. Biswell H. 1956. Ecology of california gr asslands. J Range Manage 9(1):19-24. Blank RR, Svejcar T, Riegel G. 2006. Soil attrib utes in a Sierra Neva da riparian meadow as influenced by grazing. Rangela nd Ecology & Management 59(3):321-9. 66

PAGE 75

Bock CE and W ebb B. 1984. Birds as grazing in dicator species in southeastern arizona. The Journal of Wildlife Management :1045-9. Brown JR and Archer S. 1988. Woody plant seed dispersal and the formation of herbaceous gaps: The role of domestic herbivores. PR-Texas Agricultural Experiment Station (USA) Brown JR and Archer S. 1999. Shrub invasion of grassland: Recruitment is continuous and not regulated by herbaceous bioma ss or density. Ecology 80(7):2385-96. Brown JR, Svejcar T, Brunson M, Dobrow olski J, Fredrickson E, Krueter U, Launchbaugh K, Southworth J, Thurow T. 2002. Range sites. Rangelands 24(6):712. Clary W. P. 1987. Overview of ponderosa pi ne bunchgrass ecology and wildlife habitat enhancement with emphasis on southweste rn united states. Wyoming shrublands. proceedings, 16 th wyoming shrub ecology workshop II-21. sundance, wyoming. university of wyoming, department of range management. Clary WP. 1987. Livestock effects on ripari an vegetation and adjoining streambank. Forestry Sciences Laboratory.USDA Fo rest Service. Boise, Idaho. 26p Clary WP, Shaw NL, Dudley JG, Saab VA, Kinney JW. 1996. Response of a Depleted Sagebrush Steppe Riparian System to Grazing Control and Woody Plantings.Forest Service Research Paper. Clary WP. 2000. Stubble height as a tool for management of riparian areas. J Range Manage :562-73. Clary WP and Kinney JW. 2002. Streambank and vegetation response to simulated cattle grazing. Wetlands 22(1):139-48. Collins SL, Knapp AK, Briggs JM, Blai r JM, Steinauer EM. 1998. Modulation of diversity by grazing and mowing in native tallgrass prairie. Science 280(5364):745. Cook WM, Lane KT, Foster BL, Holt RD. 2002. Is land theory, matrix effects and species richness patterns in habitat fr agments. Ecol Lett 5(5):619-23. Courtois DR, Perryman BL, Hussein HS. 2004. Vegetation change after 65 years of grazing and grazing exclusion. J Range Manage 57(6):574-82. Crampton B. 1974. Grasses in california. University of California Press. Crawley, Brown, Heard, Edwards. 1999. Invasion -resistance in experi mental grassland communities: Species richness or species identity? Ecol Lett 2(3):140-8. 67

PAGE 76

Didham RK, Tylianakis JM, Gemmell NJ, Rand TA, Ewers RM. 2007. Interactive effects of habitat modification and species invasi on on native species de cline. Trends in Ecology & Evolution, 22(9):489-96. Dull RA. 1999. Palynological evidence for 19th century grazing-induced vegetation change in the southern sierra nevada, california, USA. J Biogeogr 26(4):899-912. Dwire KA, Kauffman JB, Brookshire ENJ, Baham JE. 2004. Plant biomass and species composition along an environmental gradie nt in montane riparian meadows. Oecologia 139(2):309-17. Dwire KA, Kauffman JB, Baham JE. 2006. Plant species distribution in relation to watertable depth and soil redox potential in montane riparian meadows. Wetlands 26(1):131-46. Fleischner TL. 1994. Ecological costs of lives tock grazing in western north america. Conserv Biol 8(3):629-44. Flenniken M, McEldowney R, Frasier GW, Leininger WC, Trlica MJ. 2001. Runoff response in a montane riparian area follo wing cattle use. J Range Manage 54:567-74. Frasier G and Trlica M 2002. Sediment moveme nt and filtration in a riparian meadow following cattle use. J Range Manage: 367-73. Gelbard JL and Harrison S. 2003. Roadless habi tats as refuges for native grasslands: Interactions with soil, aspect, and grazing. Ecol Appl 13(2):404-15. Grigione MM, Beier P, Hopkins RA, Neal D, Padley WD, Schonewald CM, Johnson ML. 2002. Ecological and allometric determinants of home-range size for mountain lions (puma concolor). Anim Conserv 5(04):317-24. Grime, J. P. 1973. Competitive exclusion in herbaceous vegetation. Nature '242: 344-347. Harrison S. 1999. Native and alien species divers ity at the local and regional scales in a grazed california grassland. Oecologia 121(1):99-106. Hayes GF and Holl KD. 2003. Cattle grazing impacts on annual forbs and vegetation composition of mesic grasslands in ca lifornia. Conserv Biol 17(6):1694-702. Hewitt GB and Onsager JA. 1983. Control of grasshoppers on rangeland in the united states: A perspective. J Range Manage 36(2):202-7. Hobbs RJ and Mooney HA. 1995. Spatial and te mporal variability in california annual grassland: Results from a long-term study. Journal of Vegetation Science 6(1):43-56. 68

PAGE 77

Hofm ann L. 1983. Relationship of runoff and soil loss to ground cover of native and reclaimed grazing land. Agron J 75(4):599. Holechek JL, Valdez R, Schemnitz SD, Pi eper RD, Davis CA. 1982. Manipulation of grazing to improve or maintain wildlif e habitat. Wildl Soc Bull 10(3):204-10. Isaak DJ and Hubert WA. 2001. Production of stream habitat gradients by montane watersheds: Hypothesis tests based on spat ially explicit path an alyses. Can J Fish Aquat Sci 58(6):1089-103. Jennings MR, Hayes MP, Dept. of Fish a nd Game. 1994. Amphibian and reptile species of special concern in california. California Dept. of Fish and Game, Inland Fisheries Division. Kattelmann R. and Embury M. 1996. Ripari an areas and wetlands. Sierra nevada ecosystem project: Final report to congress. 193 p. Keeley JE, Lubin D, Fotheringham CJ. 2003. Fi re and grazing impacts on plant diversity and alien plant invasions in the southern sierra nevada. Ecol Appl 13(5):1355-74. Kie JG and Boroski BB. 1996. Cattle distribution, habitats, and diets in the sierra nevada of california. J Range Manage 49(6):482-8. Kirchner JW, Micheli L, Farrington JD. 1998. E ffects of herbaceous riparian vegetation on streambank stability. Technical Comple tion Report, Project W-872.University of California Water Resources Center Berkeley, California, USA Knapp RA and Vredenburg VT. 1996. Spawning by california golden trout: Characteristics of spawning fish, seasonal and daily timing, redd characteristics, and microhabitat preferences. Tr ans Am Fish Soc 125(4):519-31. Knapp RA, Vredenburg VT, Matthews KR. 199 8. Effects of stream channel morphology on golden trout spawning habitat and r ecruitment. Ecol Appl 8(4):1104-17. Lacey J. R. 1987. The influence of livestock grazing on weed establishment and spread. Proceedings of the montana academy of sciences (USA). Lockaby BG and Dunn BA. 1984. Camping ef fects on selected soil and vegetative properties. J Soil Water Conserv 39(3):215-6. Loft ER, Menke JW, Kie JG. 1991. Habitat shifts by mule deer: The influence of cattle grazing. The Journal of Wildlife Management :16-26. Loft ER, Menke JW, Kie JG, Bertram RC. 1987. Influence of cattle stocking rate on the structural profile of deer hiding cover. The Journal of Wildlife Management :655-64. 69

PAGE 78

Lull HW 1959. Soil compaction on forest and range lands. Magilligan FJ and McDowell PF. 1997. Stream channel adjustments following elimination of cattle grazing. J Am Water Resour Assoc 33(4). McClaran MP, Brunson MW, Huntsinger L. 2001. Future social changes and the rangeland manager. Rangelands 23(6):33-5. McEldowney RR, Flenniken M, Frasier GW, Trlica MJ, Leininger WC. 2002. Sediment movement and filtration in a riparian m eadow following cattle use. J Range Manage 55(4):367-73. Melinda A, Wheeler MA, Trlica MJ, Fras ier GW, Reeder JD. 2002. Seasonal grazing affects soil physical properties of a m ontane riparian community. J Range Manage 55(1):49-56. Milchunas DG and Lauenroth WK. 1993. Quan titative effects of grazing on vegetation and soils over a global range of environments. Ecol Monogr :328-66. Mueggler WF. 1967. Response of mountain gr assland vegetation to clipping in southwestern montana. Ecology :942-9. Mueggler WF. 1972. Influence of competition on the response of bluebunch wheatgrass to clipping. J Range Manage :88-92. Myers N, Mittermeier RA, Mittermeier CG, da Fonseca GAB, Kent J. 2000. Biodiversity hotspots for conservation prio rities. Nature 403:853-8. Niklas KJ, Midgley JJ, Rand RH. 2003. Size-dependent specie s richness: Trends within plant communities and across lat itude. Ecol Lett 6(7):631-6. ONeill KM, Olson BE, Rolston MG, Wallander R, Larson DP, Seibert CE. 2003. Effects of livestock grazing on rangeland grasshoppe r (orthoptera: Acrididae) abundance. Agriculture, Ecosystems a nd Environment 97(1-3):51-64. Olff H and Ritchie ME. 1998. Effects of herbiv ores on grassland plant diversity. Trends in Ecology & Evolution 13(7):261-5. Pattee OH. 1972. Diets of deer and cattle on the north kings deer herd summer range, fresno county, california: 1971-1972. Fresno, CA: California State Univer., Fresno. 45 p p. Pearce RA, Trlica MJ, Stednick JD, Smith JL. 1998. Sediment filtration in a montane riparian zone under simulated ra infall. J Range Manage :309-14. 70

PAGE 79

Platts W S. 1979. Relationships among stream order, fish populations, and aquatic geomorphology in an idaho river drainage. Fisheries 4(2):5-9. Platts W. 1984. Riparian system/livestock grazing interaction research in the intermountain west. Ragotzkie KE and Bailey JA. 1991. Desert mule deer use of grazed and ungrazed habitats. J Range Manage :487-90. Ratliff R. 1985. Meadows in the sierra Nevada of California: St ate of knowledge. Ries L, Fletcher RJ, Battin J, Sisk TD 2004. Ecological responses to habitat edges: Mechanisms, models, and variability explained. Annual Review of Ecology, Evolution and Systematics 35:491-522. Rothstein SI and Verner J. 1980. Range expans ion and diurnal change s in dispersion of the brown-headed cowbird in the sierra nevada. Auk :253-67. Rundel PW, Gordon DT, Parsons DJ. 1977. Montane and subalpine vegetation of the sierra nevada and cascade ranges [california]. Saab VA, Bock CE, Rich TD, Dobkin DS. 1995. Livestock grazing e ffects in western north america. Ecology and Management of Neotropical Migratory Birds: A Synthesis and Review of Critical Issues :311-53. Safford HD and Harrison SP. 2001. Grazing and s ubstrate interact to affect native vs. exotic diversity in rodeside gr asslands. Ecol Appl 11(4):1112-22. Sanders S. D. and Flett M. A. 1989. Montane riparian habitat and willow flycatchers: Threats to a sensitive environment and sp ecies. DL abell.(tech. coord.); proceedings of the california riparian systems c onference: Protection, management, and restoration for the 1990s. USDA for. serv. gen. tech. rep. PSW-110. 262 p. Seabloom EW, Borer ET, Boucher VL, Burt on RS, Cottingham KL, Goldwasser L, Gram WK, Kendall BE, Micheli F. 2003. Competiti on, seed limitation, disturbance, and reestablishment of califor nia native annual forbs. Ecol Appl 13(3):575-92. Sedgwick JA and Iko WM. 1997. Costs of brow n-headed cowbird parasitism to willow flycatchers. Studies in Avian Biology 18:167-81. Stafford MD and Valentine BE. 1985. A prelim inary report on the biology of the willow flycatcher in the central sierra nevada California-Nevada Wildlife Trans :66-7. Stohlgren TJ, DeBenedetti SH, Parsons DJ. 1989. Effects of herbage removal on productivity of selected hi gh-sierra meadow community types. Environ Manage 13(4):485-91. 71

PAGE 80

Stolgren TJ, Chong GW, Schell LD, Ri mar KA, Otsuki Y, Lee M, Kalkhan MA, Villa CA. 2002. Assessing vulnerability to invasi on by nonnative plant species at multiple spatial scales. Envir on Manage 29(4):566-77. Storer TI, Usinger RL, Lukas D. 2004. Sierra ne vada natural history. 1st ed. University of California Press. 439 p. Strand S. 1979. Pack stock management in the high sierra. Taylor DM. 1986. Effects of cattle grazing on passer ine birds nesting in riparian habitat. J Range Manage 39(3):254-8. Tilman D. 1997. Community invasibilit y, recruitment limitation, and grassland biodiversity. Ecology 78(1):81-92. Tilman D and Downing JA. 1996. Biodiversity and stability in grasslands. Ecosystem Management: Selected Readings. U.S. Fish and Wildlife Service Website. 2009. California golden trou t. http://ecos.fws. gov/speciesprofile/pro file/speciesprofile.action?spcode=E083 USDA Forest Service Website. 2009. http ://www.fs.fed.us/psw/programs/snrc/bio diversity /aquatic_ecosystems_sub3/ reintro_declining_amphibia.shtml. Vale TR. 1981. Tree invasion of montane meadows in oregon. Am Midl Nat :61-9. Vale TR. 1987. Vegetation change and park purpo ses in the high eleva tions of Yosemite national park, california. Ann Assoc Am Geogr :1-18. Vale TR. 1977. Forest changes in the warner mountains, california. Ann Assoc Am Geogr 67(1):28-45. Van Wieren SE. 1998. Effects of large herb ivores upon the animal community. Grazing and Conservation Management :185. Van Woudenberg A. M. 1999. Grazing impact s on the biodiversity of riparian ecosystems. At risk: Proceedings of a c onference on the biology and management of species and habitats at risk. BC min. e nviron., lands and parks, victoria, BC. Vankat JL and Major J. 1978. Vegetation change s in sequoia national park, california. J Biogeogr :377-402. Vavra M, Laycock WA, Pieper RD. 1994. Ecologi cal implications of livestock herbivory in the west. 72

PAGE 81

73 Verner J and Ritter LV. 1983. Current status of the brown-headed cowbird in the sierra national forest. Auk :355-68. Weaver W and Shannon CE. 1959. The mathematical theory of communication. University of Illinois Press Urbana. Wilkerson RL and Siegel RB. 2001. Establishing A southern sierra meadows important bird area: Results from meadow surveys at stanislaus, sierra and sequoia national forests, and yosemite and Sequoia/Kings canyon national parks. The Institute for Bird Populations, Pt.Reyes Station, CA Williamson J and Harrison S. 2002. Biotic and ab iotic limits to the spread of exotic revegitation species. Ecol Appl 12(1):40-51. Wood SH. 1975. Holocene stratig raphy and chronology of m ountain meadows, sierra nevada, california. California Institute of Technology. Jennings M. R. 1996. Status of amphibians. Sierra nevada ecosystem project: Final report to congress. 921 p. Woolfenden WB. 2003. A 180,000-year pollen record from owens lake, CA: Terrestrial vegetation change on orbital sc ales. Quatern Res 59(3):430-44.