Causes and consequences of woody plant encroachment into western North American grasslands

Journal of Environmental Management 90 (2009) 2931–2942

Contents lists available at ScienceDirect

Journal of Environmental Management journal homepage: www.elsevier.com/locate/jenvman

Review

Causes and consequences of woody plant encroachment into western North American grasslands O.W. Van Auken* Department of Biology, University of Texas at San Antonio, San Antonio, TX 78249-0062, USA

a r t i c l e i n f o

a b s t r a c t

Article history: Received 16 October 2008 Received in revised form 18 April 2009 Accepted 30 April 2009 Available online 5 June 2009

As woody plants encroach into grasslands, grass biomass, density and cover decline as wood plant biomass, density and cover increase. There is also a shift in location of the biomass from mostly belowground in the grasslands to aboveground in the woodlands. In addition, species richness and diversity change as herbaceous species are replaced by woody species. This is not a new phenomenon, but has been going on continually as the climate of the Planet has changed. However, in the past 160 years the changes have been unparalleled. The process is encroachment not invasion because woody species that have been increasing in density are native species and have been present in these communities for thousands of years. These indigenous or native woody species have increased in density, cover and biomass because of changes in one or more abiotic or biotic factors or conditions. Woody species that have increased in density and cover are not the cause of the encroachment, but the result of changes of other factors. Globally, the orbit of the Earth is becoming more circular and less elliptical, causing moderation of the climate. Additional global climate changing factors including elevated levels of CO2 and parallel increases in temperature are background factors and probably not the principal causes directing the current wave of encroachment. There is probably not a single reason for encroachment, but a combination of factors that are difficult to disentangle. The prime cause of the current and recent encroachment appears to be high and constant levels of grass herbivory by domestic animals. This herbivory reduces fine fuel with a concomitant reduction in fire frequency or in some cases a complete elimination of fire from these communities. Conditions would now favor the woody plants over the grasses. Reduced grass competition, woody plant seed dispersal and changes in animal populations seem to modify the rate of encroachment rather than being the cause. High concentrations of atmospheric CO2 are not required to explain current woody plant encroachment. Changes in these grassland communities will continue into the future but the specifics are difficult to predict. Density, cover and species composition will fluctuate and will probably continue to change. Increased levels of anthropogenic soil nitrogen suggest replacement of many legumes by other woody species. Modification and perhaps reversal of the changes in these former grassland communities will be an arduous, continuing and perhaps impossible management task. Ó 2009 Published by Elsevier Ltd.

Keywords: Arid grassland Brush encroachment Brush invasion Desert grassland Desertification Drylands Encroachment Grasslands Rangelands Semiarid grassland Woody plant encroachment Woody plant invasion

1. Introduction Encroachment is the increase in density, cover and biomass of indigenous woody or shrubby plants in various grasslands, especially arid and semiarid grasslands (Fig. 1) (Van Auken, 2000). Many use encroachment and invasion interchangeably, although invasion would be coming from another continent or a great distance. There are many paired, temporal photographs that show that encroachment has occurred, but they do not show the cause of the

* Tel.: þ1 210 458 5489; fax: þ1 210 458 5658. E-mail address: [email protected] 0301-4797/$ – see front matter Ó 2009 Published by Elsevier Ltd. doi:10.1016/j.jenvman.2009.04.023

encroachment. The best evidence demonstrating encroachment is found in studies that examined a specific grassland area and changes in area covered or density of one or more woody species in time (Fig. 2) (Knapp et al., 2008a). In some cases, canopy closure can be completed within 40 years. However, most encroachment occurred well before scientists began looking at this phenomenon in a systematic way. Because woody plants can reach an appreciable age, another type of study is possible. One can examine date of woodland community establishment retrospectively using tree-ring chronologies or dendrochronology. A large number of communities are aged and the ages of the communities are examined as a function of time. When this was done, a large, rapid increase in the number of

2932

O.W. Van Auken / Journal of Environmental Management 90 (2009) 2931–2942

Fig. 3. The number of newly established Juniperus occidentalis communities per decade for the past 400 years. There were 801 communities examined in former northern California grasslands (modified from Miller et al., 2005).

Fig. 1. Encroachment of Juniperus plants into semiarid grassland west of Albuquerque, New Mexico. View is of Enchanted Mesa taken from Acoma Pueblo. Picture on the top was taken by W.H. Jackson in 1899 and the one on the bottom by H.E. Malde in 1977 (Allen et al., 2002).

recently established woodland communities was found (Miller et al., 2005; Miller et al., 2008) (Fig. 3). The increase in the number of new woodland communities started in the mid to late 1800s and continued through most of the 1900s. It is important to understand this phenomena because arid and semiarid lands, or drylands, cover about 41% of the terrestrial surface of the Earth with approximately 2.4 billion people living in these habitats (GLP, 2005; MEA, 2005). This area includes mostly rangelands or pastoral lands with very limited dryland farming (Campbell et al., 1997; Morgan et al., 2007). A considerable portion

Fig. 2. Increase in mean cover ( 1 SE) of Juniperus virginiana in northeastern Kansas. Area began as grassland in 1956 and 40 years later it was forest community with 95þ% cover (modified from Knapp et al., 2008a).

of this area has undergone encroachment of woody plants and severe degradation or desertification (10–20% as a medium confidence estimate) (Reynolds et al., 2007). Desertification, is considered land degradation in arid, semiarid and dry sub-humid areas resulting from climatic factors and human activity (Reynolds et al., 2007). It is very difficult to ascribe a single factor as the cause of the desertification. Actually, there appear to be several interacting factors that are involved in promoting desertification and causing the widespread encroachment of woody plants into these arid and semiarid grasslands (Van Auken, 2000). One factor concerns the Earth’s orbit which is becoming more circular and less elliptical (Milankovitch, 1998; Mackenzie, 2003), which is a long-term, cyclic change and background to anthropogenic changes occurring today. A second type of factor would include atmospheric CO2 levels, rainfall, temperature, soil depth, nutrient levels, fire intensity and frequency and amount and kind of herbivory. It appears that several of these factors are at work causing the recent encroachment of woody plants into grasslands in southwestern North America as well as in other arid and semiarid parts of the World. As the Earth’s orbit becomes more circular, the climate is moderated because of a change in the amount of solar energy that reaches the Earth. Orbital changes occur over thousands of years and are related to the total summer radiation received in northern latitudes where ice sheets have formed in the past (Imbrie and Imbrie, 1980). These cycles have been connected to changes in the climate of the Earth over the past 420,000 years (Petit et al., 1999) and linked to glacial recession and northern migration of North American plant communities over the last 11,000–12,500 years (Delcourt et al., 1983; Betancourt et al., 1990; Van Devender, 1995). The latest phase of the Milankovitch eccentricity cycle, the warmer part of the cycle is called the Holocene Epoch and began between 11,000 and 12,500 years ago. This is basically the time commencing with the end of the last glaciation. However, plant and animal communities and populations have been changing at various rates since the earliest times (Berner, 2005). Deglaciation during the Holocene has caused major fluctuations in extent and composition of plant communities throughout the world and is well documented, especially in North America (Delcourt et al., 1983; Betancourt et al., 1990; Miller and Wigand, 1994; Van Devender, 1995). The causes of the deglaciation and migration of plant communities from low to higher latitudes and from low to higher altitudes seem to be increased mean annual temperature

O.W. Van Auken / Journal of Environmental Management 90 (2009) 2931–2942

which is coupled to higher levels of atmospheric CO2. The changes in all of these factors were gradual compared to changes in many plant communities seen in the past 160 years. In the eastern United States, where areas of deciduous forest are presently found there were spruce, fir and pine Boreal forest which retreated to the north (Delcourt et al., 1983). Subalpine woodlands were found in areas now considered pinion–juniper woodlands in the Great Basin (Betancourt et al., 1990; Miller and Wigand, 1994; Van Devender, 1995). Spruce forests or open spruce, pine, birch parkland was found in areas of tall-grass prairie in Kansas and in areas of Juniperus grassland or savanna in west central New Mexico. Pine parkland was originally present in what is now short grass prairie in western Texas and desert grassland in southeastern Arizona (Van Devender, 1995). As indicated, these changes occurred over the past 11,000–12,500 years and were relatively gradual spatial and temporal changes with considerable flux in species composition, density and area covered. 1.1. The recent past Recent changes in distribution and density of plant populations and communities (the past 160 years) have been rapid and unparalleled (Pimm et al., 1995). These recent changes in woody plant populations associated with grasslands or savannas do not appear to be caused directly by deglaciation or to be considered invasions as is often implied. They are trends associated with woody species found in adjacent communities that have been in existence for a considerable time (Miller and Wigand, 1994). In the arid grassland communities at the lower elevations of southwestern North America, the woody and succulent species usually associated with the dry lands of the Chihuahuan or Sonoran Deserts have increased in density (Buffington and Herbel, 1965; Hastings and Turner, 1965; Browning et al., 2008). At the upper reaches of these grasslands usually at higher elevations, various species of Juniperus, previously restricted to rocky outcrops, steep slopes and shallow soils, have spread down slope into semiarid grasslands (Wells, 1965; Miller and Wigand, 1994; McPherson, 1997; Miller et al., 2008). Similar increases in woody plants have occurred in other grasslands in other places including the great plains of central North America (McKinley and Blair, 2008; McKinley et al., 2008a; Knapp et al., 2008b). Changes in density of woody plants in the higher elevation portions of these grasslands have been attributed by many to climate change (Buffington and Herbel, 1965; Hastings and Turner, 1965), but in these grasslands, warming suggests that Juniperus populations would move north and up in elevation rather that down slope like they have moved (Miller and Wigand, 1994). Recent measurements of temperature and precipitation at low elevations in southwestern North America and modeled predictions of future climatic trends show a warmer drier climate, especially during summer (Seager et al., 2007; Solomon et al., 2009). However, in spite of these studies, links between recent changes in climate and recent changes in vegetation patterns in this area are weak (Bahre and Shelton, 1993). Large-scale droughts have been related to widespread die-offs and decrease in pinion pine populations in some of these Pinion-Juniper woodlands (Breshears et al., 2005; Breshears, 2008), but encroachment of Juniperus plants into associated grasslands seems to be caused by management decisions related to grazing caused reduced grass biomass and a concomitant reduction in fire frequency (McPherson et al., 1988). There are many examples of species that have moved or encroached into various grasslands (Archer et al., 1995; Van Auken, 2000; Roques et al., 2001; Silva et al., 2001; Hoch et al., 2002; Geist and Lambin, 2004; Wessman et al., 2004; Briggs et al., 2005; Fensham et al., 2005; Havstad et al., 2006; Peters et al., 2006; Ryniker

2933

et al., 2006; McKinley and Blair, 2008; Van Auken and Smeins, 2008; Knapp et al., 2008b), but woodland populations of Quercus emoryi (Emory oak) in parts of Arizona and Northern Mexico have been relatively stable for many years (Weltzin and McPherson, 1999). Most of the recent changes in these semiarid grasslands seem to be directly or indirectly caused by anthropogenic factors or management decisions. The process is called desertification, shrub invasion, woody plant invasion and brush or woody plant encroachment and has occurred throughout the world (Archer et al., 1995; Van Auken, 2000; Roques et al., 2001; Silva et al., 2001; Hoch et al., 2002; Geist and Lambin, 2004; Wessman et al., 2004; Briggs et al., 2005; Fensham et al., 2005; Havstad et al., 2006; Peters et al., 2006; McKinley and Blair, 2008; McKinley et al., 2008a; Van Auken and Smeins, 2008; Knapp et al., 2008b). These encroached ecosystems are difficult to study because of their biphasic nature, an open grassland phase and a closed woodland phase. Various densities of woody and herbaceous plants are found together and one type of plant should not be examined independently of the other. In arid and semiarid regions there appears to be a continuum from grassland on one end to woodland or forest on the other end of the continuum (House et al., 2003). As one proceeds from the grassland into the woodland, the grass cover and biomass decrease as the woody plant cover and biomass increase (Fig. 4) (Wayne and Van Auken, 2008). In addition, species richness (number of species) decreases as one proceeds from the grassland into the associated encroached woodland (Fig. 5) (Knapp et al., 2008a). Other ecosystem factors change as well including productivity, location of biomass (mostly belowground in the grassland to mostly above ground in the woodland), type of biomass (herbaceous to woody), and various soil processes (McKinley and Blair, 2008; Throop and Archer, 2008). The communities between the grasslands and woodlands are savannas with high variation in woody plant density. Encroachment of woody plants is occurring all along this continuum converting grassland and savanna to woodland and forest. The level of productivity, structure and time involved in the conversions of these diverse ecosystems will depend on the interaction of rainfall, temperature, soils, fire, herbivory and anthropogenic factors. Independent manipulation and control of these factors are at best difficult to accomplish (House et al., 2003). 1.2. Region of focus This report focuses mainly on the dry lands, the arid and semiarid grasslands or desert grasslands in the basins and valleys of

Fig. 4. Changes in mean cover ( 1 SE) of grasses and woody plants proceeding from the grassland into adjacent Juniperus ashei woodland in central Texas, USA (modified from Wayne and Van Auken, 2008).

2934

O.W. Van Auken / Journal of Environmental Management 90 (2009) 2931–2942

Fig. 5. Decrease in herbaceous species richness with an increase in Juniperus virginiana density in northeast Kansas (from Knapp et al., 2008a).

southwestern North America as well as the outwash plains (bajadas) in this same region (McClaran and Van Devender, 1995; House et al., 2003). Encroachment in this area and the factors that seem to control the process are similar in various ecological regions and grasslands throughout North America (McKinley and Blair, 2008; McKinley et al., 2008a), South America (Soriano, 1979; Cabral et al., 2003), southern Africa (Cole, 1986; MacDonald, 1989), Asia (Misra, 1983; Sharma and Dakshini, 1991) and Australia (Perry, 1970; van Klinken and Campbell, 2001). Low elevation (1100 and 2500 m) semiarid grasslands of southwestern North America are extensive, but discontinuous (McClaran and Van Devender, 1995). They extend from northern Texas, across southern New Mexico, to central Arizona (approximately 35 N, 98 W to 31 N, 111 W). In Mexico, they occur west of Veracruz to northeast of Colima (approximately 18 N, 96 W to 19 N, 104 W). There is considerable climatic variability across this vast but discontinuous area. Precipitation in the north maybe 230 mm per year in places and as much as 600 mm per year at higher elevations in the south. Rainfall is usually summer rain with 50–90% falling between May and October. Mean annual temperature ranges from 13 to 16  C with fewer than 75 days with freezing temperatures and the evaporation rates are high. Aboveground net primary production is as low as 43 g m2 yr1 with total plant production between 250 and 350 g m2 yr1. These are the lowest values for all of the North American grasslands, probably because of low and variable rainfall, high evapotranspiration and shallow soil. The soils of this diverse area are mostly shallow and usually Aridisols, Mollisols or Entisols (Bailey, 1978). Included are sandy outwash materials with little horizon development to old, deep soils with well-developed profiles. Deeper soils have well-developed calcic horizons, but all of the soils have less organic material than other grasslands. Soil properties of these habitats including the distribution of water and the capacity to hold water determine to a large extent the kind of plants that are present (McClaran and Van Devender, 1995).

1.3. Recent changes and introductions Reports of increased density of woody plants in the arid and semiarid grasslands of southwestern North America date to the mid to late 1800s. This encroachment in Arizona is reasonably well documented, but still mostly anecdotal. It appears to be linked or coupled to increased cattle ranching in this area in the 1870s and

the concomitant reduction in fire frequency about the same time (Bahre, 1991; Bahre and Shelton, 1993). Additional evidence of dramatic changes in this area during the late 1800s and early 1900s has recently been presented. Dust deposition increased in parts of southwestern North America by 500% above the late Holocene average (Neff et al., 2008). This increase is linked to human activity and was probably caused by rapid expansion of livestock grazing at this time. This would be associated with reduced grass cover, reduced light fuel levels and a concomitant reduction in fire frequency, conditions that would favor the establishment, encroachment and growth of woody plants over the grasses. There is some documentation of the same phenomena in New Mexico and Texas as well, but possibly starting sooner (Humphrey, 1958; Dick-Peddie, 1993; Archer, 1994). Nonetheless, increased woody plant density occurred before the major influx of CO2 into the atmosphere but the increased density was associated with increasing temperature, which was a background factor. During this time or later, several species of woody plants were introduced into southwestern North America and became established. However, none are major invading species in this area. Euryops multifidus (resin bush), was introduced from South Africa (McClaran and Van Devender, 1995), several species of Tamarix (Salt cedars or Tamarisk) were introduced from the Middle East or Africa (Correll and Johnston, 1970), and Elaeagnus angustifolia (Russian olive) from Asia (Humphrey, 1958; Stromberg et al., 1997). Truly invasive herbaceous species came from various places in the Middle East and Africa (Mooney and Drake, 1986; Brock et al., 1997). Heavily grazed, disturbed habitats are more likely to be invaded than similar non-disturbed habitats. These species are usually annuals (House et al.) from the families Poaceae, Asteraceae (sunflower) and Brassicaceae (mustard). Approximately 10% of the local floras from southwestern North America are established, nonnative species and their rate of arrival is unknown (Enserink, 1999; Kaiser, 1999; Malakoff, 1999). 1.4. Area encroached The area of grassland in southwestern North America that has been converted to savanna, shrubland or woodland is estimated as high as 60 million ha (Humphrey, 1958; Grover and Musick, 1990). This is more than the total area of the semiarid grasslands of southwestern North America (Lauenroth, 1979). Estimations have been difficulty to make because more that one species is found in the same area as part of a community (Grover and Musick, 1990). Some of the encroaching species are found in areas outside the traditional semiarid grasslands with no habitat or area differentiation. Furthermore, density and cover of woody plants are highly variable with some areas having high density or cover, others with moderate density and some with low density. Consequently, some consider these areas grasslands, while others consider them savannas, shrublands or woodlands. The conclusion, species composition, cover and density or structure of most if not all of these southwestern North American arid and semiarid grasslands have changed because of the encroachment of one or more woody species. This encroachment in North American grasslands is not limited to the arid and semiarid southwest and some suggest that between 220 and 330 million ha of total grassland have been or are being encroached (Knapp et al., 2008b). The main genus of woody plant involved in this encroachment in southwestern North America is Prosopis (mesquite, Prosopis glandulosa, and probably Prosopis velutina, Prosopis torreyana or Prosopis juliflora). They seem to be the dominant woody plants on over 38 million ha of what has been considered semiarid southwestern grasslands. Larrea tridentata (creosotebush) is the dominant shrub on over 19 million ha of similar former grassland

O.W. Van Auken / Journal of Environmental Management 90 (2009) 2931–2942

habitats. There are additional species of shrubs, sub-shrubs, small trees and succulents that have encroached into these and other grasslands including various species of Acacia, Cercidium (Paloverde or retama), Yucca, Flourensia (tar bush), Haplopappus (goldenweed or camphor weed), Opuntia (prickley pears and chollas), Gutierrezia (broomweeds and snakeweeds), Juniperus (juniper or mountain cedar), and Quercus (oak) (Humphrey, 1958; Grover and Musick, 1990; Schmutz et al., 1991; Dick-Peddie, 1993). These additional species are locally important, but individually do not cover the same area as do Prosopis and Larrea at low elevation. In slightly cooler higher elevation grasslands or more northern or eastern grasslands Juniperus is one of the major encroaching species (Miller and Wigand, 1994; Miller et al., 2005; Miller et al., 2008; Van Auken and Smeins, 2008; Knapp et al., 2008b). 1.5. Potential causes of encroachment Considerable discussion and debate have centered on the factors that cause woody plant encroachment in the dry lands of the World (Schlessinger et al., 1990; Archer et al., 1995; Van Auken, 2000; Silva et al., 2001; Hoch et al., 2002; Geist and Lambin, 2004; Fensham et al., 2005; Havstad et al., 2006; Peters et al., 2006; Knapp et al., 2008a; Knapp et al., 2008b). These debates have included changes in the arid and semiarid grasslands of southwestern North America. Usual reasons cited are global climate change, chronic high levels of herbivory, change in fire frequency, changes in grass competitive ability, spread of seed by livestock, small animal populations, elevated levels of CO2, and combinations of these factors. Most of the changes in density of woody plants in southwestern grasslands have occurred in the past 160 years and have been associated with the introduction of cattle and cattle grazing systems (Bahre, 1991; Bahre and Shelton, 1993; Bartolome, 1993; Archer et al., 1995). Herbivory is tolerated by grasses without noticeable changes in productivity, biomass, growth or reproduction, but only a small amount. With higher levels of herbivory, there is a depression or reduction of all of these factors (Harper, 1977; Belsky, 1986; Gardener et al., 1990; Louda et al., 1990; Heitschmidt and Stuth, 1991; Archer, 1994; Begon et al., 2006). Reduction of aboveground grass biomass by herbivory can be rapid, but re-growth can also be swift. Small scale changes in grass biomass or productivity over a long time are not as easy to see because the aboveground parts are replaced annually and relatively quickly by mobilization of belowground stored resources with considerable year to year variation. Once soil resources are used up, aboveground grass biomass and productivity will decline. Thus, high levels of longterm herbivory will reduce both aboveground and belowground grass biomass, reducing the grass’s ability to re-grow, to replace itself and to complete. When this occurs with a concomitant reduction in fire frequency, the woody plants are favored over the grasses and their establishment and growth is promoted. Direct stimulation, compensation or overcompensation of aboveground herbaceous plant growth by herbivory and benefits of herbivory to plants seems minimal but is reported and these topics have received considerable attention in the literature (Briske, 1993; Dyer et al., 1993; McNaughton, 1993; Painter and Belsky, 1993; Belsky, 1996; Belsky and Blumenthal, 1997; Begon et al., 2006). Grasses do not benefit directly from herbivory (Begon et al., 2006). However, there may be some growth stimulation, but the initial source of the carbon or carbohydrates used for re-growth is most likely reallocation of belowground biomass. Thus, chronic high levels of herbivory seem to negatively affect the growth of grasses and other herbaceous plants and seem to be the dominant reason or the driving force for increases in woody plant density in the arid and semiarid grasslands of southwestern North America and other places (Bahre, 1991; Bartolome, 1993;

2935

Noy-Meir, 1993; Patten, 1993; Archer et al., 1995; Bahre, 1995). Although high populations of domestic cattle seem to be the primary factor in the conversion of arid and semiarid grasslands into shrublands, savannas and woodlands, the mechanisms involved are still not well understood and the rates, dynamics, patterns and successional processes are still being defined. Consequently, the specific mechanism allowing woody plant encroachment into the dry lands of southwestern North America and the World seem to be complex.

1.6. Global change Throughout the Pleistocene, most of the past 2 million years, the climate of the Earth was cooler than today (McDowell et al., 1995; Van Devender, 1995; Martin, 1999). Possibly 15–20 glacial periods occurred with associated interglacials (Imbrie and Imbrie, 1979). The interglacials were 10,000–20,000 years long and the glacials about ten times longer. The present interglacial started about 12,500 years ago at the beginning of the Holocene Epoch and is projected to last much longer than most of the previous interglacials (Berger and Loutre, 2002). As temperatures warmed, the glaciers melted and plant communities around the world moved and changed their distributions as indicated previously. In the past, plant communities migrated as the climate warmed or cooled without the influence of man or his animals. During the current warming trend (Holocene), in what are now the semiarid grasslands or desert grasslands of southwestern North America, pine parkland and juniper woodland or savanna moved mostly in a northern direction or up in elevation. The semiarid grasslands of the American southwest were lower in elevation and more to the south, before migrating to their approximate current location. Chihuahuan and Sonoran Desert shrublands were lower in elevation and more to the south. With warming they migrated to there approximate, current locations. Changes in distributions of populations of plants and animals in the American southwest over the past 11,000–12,500 years have been developed from pollen records and fossil packrat middens (Betancourt et al., 1990; Miller and Wigand, 1994; Van Devender, 1995; Martin, 1999). Radiocarbon dating and oxygen isotope data from deep sea cores have been linked to develop a single time series to represent global changes in this and other areas (McDowell et al., 1995). This current warming trend is expected to continue with a prolonged interglacial projected for possibly another 50,000 years (Berger and Loutre, 2002). However, it has been difficult to link recent vegetation changes to climatic changes over the past 160 years because of interacting factors and the relatively short time period. Modeled predictions of current and future climatic trends in southwestern North America show a warmer drier climate, especially during summer (Seager et al., 2007; Solomon et al., 2009). However, links between changing climate since the 1870s and recent shrub or woody plant encroachment in the semiarid grasslands are weak (Bahre and Shelton, 1993). Recent drought has caused increased but unequal woody plant mortality in this area (Breshears et al., 2005), but not in all communities. The severity and frequency of similar droughts are expected to increase as the climate continues to get warmer and drier. However, the unevenness of encroachment, especially for Prosopis and Juniperus, and dramatic differences in density in adjacent, fenced, edaphically similar areas, would seem to rule out large-scale climatic influences as the major cause of recent increased woody plant density (Fig. 6) (Bahre and Shelton, 1993; Archer et al., 1995). Management decisions seem to be more important in determining the type of community, the kind of species present and the type of biomass that would be present (herbaceous or woody) (Fig. 7).

2936

O.W. Van Auken / Journal of Environmental Management 90 (2009) 2931–2942

Fig. 6. Aerial photograph of a fence line showing the effect of management differences with higher woody plant density and cover on the right and lower woody plant density and cover on the left.

1.7. Elevated levels of atmospheric CO2 Considerable evidence is available to demonstrate that a number of C4 taxa replaced some C3 taxa during past glacial periods with a reversal of the trend during subsequent interglacials (Ehleringer, 2005). This would correspond to times with low atmospheric CO2 followed by periods of higher levels of CO2. The current trend of increasing atmospheric CO2 has been going on through the previous 11,000–12,500 years of the Holocene, since the end of the last glacial maximum. Consequently, increased levels of atmospheric CO2 have been proposed by some as the cause of recent C3 shrub and woody plant encroachment into grasslands throughout North America including the semiarid grasslands in the southwest (Mayeux et al., 1991; Idso, 1992; Polley et al., 1992; Johnson et al., 1993). However, the replacement species in previous times were probably C4 and C3 grasses, not C3 woody plants. The evidence does not purport to show that the shift was from C4 grasses to C3 woody plants (Ehleringer, 2005). There is however some recent evidence to link greater woody plant density to greater growth at higher levels of atmospheric CO2 (Morgan et al., 2007). This hypothesis is very attractive and could account for the synchronous, widespread encroachment of woody plants into semiarid grasslands and savannas in the past 160 years; however, proof is equivocal. The elevated CO2 hypothesis is based on observations that most woody plants have the C3 photosynthetic pathway and many of the grasses that are being replaced in southwestern grasslands have the C4 photosynthetic pathway, however some of these grasses being replaced have the C3 photosynthetic pathway. The C3 photosynthetic pathway is advantageous at higher levels of CO2 (Ehleringer,

2005) and is present in all of the cool season or C3 grasses. There are some difficulties with this replacement hypothesis (Archer et al., 1995). Many C3 and C4 species have similar quantum yields, photosynthesis rates and water-use efficiencies. Many C4 grasses are more responsive to increased levels of CO2 than initially thought. Replacement of C3 grasses by encroachment of C3 woody shrubs in the warm deserts or cold deserts is not explained by the elevated CO2 hypothesis. In southwestern North American semiarid grasslands, C3 woody plants are replacing both the C4 grasses and C3 grasses. In addition, the C3 grasses are not replacing C4 grasses. In areas with similar soils, fences preventing the movement of domestic herbivores or having different management regimes reduce or prevent the encroachment of C3 woody plants (Fig. 7). Replacement of populations of C4 grasses with C3 grasses which might be expected has not happened. If one examines available data and the timing of greatest woody plant encroachment, it does not match the highest levels of atmospheric CO2. There is a temporal disparity between these two factors. Highest levels of atmospheric CO2 (today) followed the greatest extent of woody plant encroachment (1850–1870). Populations of woody plants and grasses have shifted during the Holocene, and the shifts seem to be explained by elevated temperatures driven by higher levels of atmospheric CO2. But, the major shifts or encroachment of woody plant populations that peaked in the late 1800s does not correspond to the highest levels of atmospheric CO2 (today). In addition, not all elevated CO2 studies have shown a fertilizer effect; suggesting other limitations or constraints on the grasses that are being replaced. Constraints could be limitations of soil resources such as nitrogen, phosphate, water or a combination of these or possibly other factors. Consequently, the CO2 enrichment hypothesis does not seem to explain the encroachment of woody plants into the southwestern semiarid grasslands or other grasslands. Another factor or factors seem to be suggested. In spite of temporal constraints, various studies have suggested reasons that the C3 woody plants should replace the C4 grasses in the arid and semiarid grasslands of southwestern North America (Shaw et al., 2005). These arid and semiarid grasslands should respond positively to elevated levels of CO2 because gas exchange and other plant responses seems to be limited primarily by available water (Mooney et al., 1991). Many C3 species are stimulated to higher growth rates than many C4 species (Long, 1991; Poorter and Navas, 2003). In addition, the consequence of long-term, high levels of CO2 exposure are probably dependent on temperature, precipitation, disturbances, especially herbivory and may not have a single outcome and will probably be complex (Shaw et al., 2005). There are numerous proposed and possibly expected changes in these arid and semiarid ecosystems in a future high CO2 environment, but outcomes are basically unknown and empirical evidence is limited or nonexistent. Other factors or multiple factors will have to be considered.

Fig. 7. View along a fire-break on the Konza prairie, northeast Kansas (39 050 N and 96 350 W) showing effects of different management treatments on tall-grass prairie. The right side of the fire-break (center) has been burned frequently (every 2 years) and on the left side, infrequently (less than once every 10 years). Photos by D.C. McKinley.

O.W. Van Auken / Journal of Environmental Management 90 (2009) 2931–2942

2937

1.8. Competition Positive and negative interactions seem to be important determinants of community structure and function but this competition is a constantly debated topic (Van Auken, 2000). Plant growth in arid and semiarid communities as in other communities is controlled to some extent by neighbors and can be drastically influenced by herbivory. Encroachment of woody plants into semiarid grasslands is modified by competition or slowed by the presence of grasses and belowground competition seems to be critical (Bush and Van Auken, 1986; Van Auken and Bush, 1989; Bush and Van Auken, 1995; Van Auken and Bush, 1997). Root biomass, surface area and structure are certainly important as are other root characteristic and certain environmental conditions. When woody plant seeds germinate in grasslands and shortly after, the grasses have the competitive advantage, but the interaction is reversed once the woody plants reach a certain size (Bush and Van Auken, 1986; Van Auken and Bush, 1989; Bush and Van Auken, 1995; Van Auken and Bush, 1997). This would be true for both the C3 and C4 grasses. The roots of the woody plants seem to have to be below the zone of the grasses roots (Walker, 1971; Mlambo et al., 2005) and the stem and other aboveground parts have to be above the leaf zone of the grasses where light levels are higher (Bush and Van Auken, 1995; Van Auken and Bush, 1997). In addition, the woody plant stem and bark have to be thick enough to resist the temperatures associated with grassland fires and the leaves and buds have to be high enough above the fire not to be effected by the elevated air temperatures of the passing flames (Collins and Wallace, 1990). Considerable evidence has been presented to demonstrate the interactions of grasses and woody plants, especially for mesquite and acacia (P. glandulosa and Acacia smallii). Biomass of the woody plants is reduced when grown with various grasses in both greenhouse and field studies (Bush and Van Auken, 1986; Van Auken and Bush, 1989; Bush and Van Auken, 1995; Van Auken and Bush, 1997). Exceptions occur, probably caused by lower levels of herbaceous biomass or site specific factors. Woody plant size and biomass are reduced in the presence of the grasses. If the grass biomass is removed by high levels of constant domestic herbivory, the woody plant grows as if the grass is not present (Van Auken and Bush, 1989). There are some specific difficulties with the grass competition theory. Woody plant growth is suppressed but total mortality usually does not occur. A few of the woody plant seedlings will finally escape the grass zone of suppression and survive to become shrubs or small trees, which may convert the grassland into a savanna, shrubland or woodland. Time is a critical factor, with only a few survivors per year over 100 years, a grassland would be converted to a savanna and then a shrubland or woodland. Negative effects of grass competition reduce the rate of change of grassland to woodland or shrubland but do not prevent the eventual change. It seem that another factor in addition to grass competition is required to maintain grassland communities as grasslands and prevent the encroachment of woody plants and that factor seems to be fire.

Fig. 8. Juniperus ashei encroachment into central Texas mid-grass prairie (photo by the author).

2007). Recent modeling studies have produced similar findings, heavy grazing or low fire frequency results in encroachment followed by increased woody plant density (Fuhlendorf et al., 2008). There are interactions that occur between fires and other factors including topography, soil type, number and kind of herbivores and amount of light-fluffy fuel. These factors determine the nature, density and location of woody plants in a given landscape (Humphrey, 1958; Collins and Wallace, 1990; McPherson, 1995). In addition, fire frequency and intensity are linked to climatic patterns and conditions. Forest communities in the southwestern United States have fire patterns that are determined by rainfall and temperature that are in turn determined by Southern Oscillations ˜ a, low phase-El Nin ˜ o) (Swetnam and Betancourt, (high phase-La Nin 1990). In this area of western North America, large fires usually ˜ a) and smaller fires follow wet occur after dry springs (La Nin ˜ o). This pattern is probably true for the lower springs (El Nin elevation semiarid grasslands of this area as well. Early evidence of rangeland or grassland fires in southwestern North America is relatively sparse and not from the usual sources. Historical evidence comes from newspaper reports, from the earliest travelers, and from early settlers (Humphrey, 1958; Inglis, 1964; Bahre and Shelton, 1993). But some do not think there was ever enough fuel in these grasslands to carry an extensive fire or

1.9. Fire The importance of periodic fire in grasslands cannot be overestimated. If intermittent fires do not occur in grasslands, they will be converted to savannas, shrublands or woodlands (Figs. 7–9). Fire frequency is variable, but there is general agreement that recurring fires are required to control or reduce the establishment, density and growth of woody plants in most if not all grasslands (Collins and Wallace, 1990; Scholes and Walker, 1993; Simmons et al.,

Fig. 9. Prosopis glandulosa encroachment into central Texas mid-grass prairie (photo by the author).

2938

O.W. Van Auken / Journal of Environmental Management 90 (2009) 2931–2942

don’t agree with the evidence (Buffington and Herbel, 1965; Hastings and Turner, 1965; Dick-Peddie, 1993). Changes in the composition of the low elevation grasslands in many parts of southwestern North America seemed to take place after the beginning of large-scale cattle ranching and fire exclusion in the 1850s–1870s (Humphrey, 1958; Bahre, 1991; Bahre, 1995; Neff et al., 2008). Wildfires in these low elevation communities today are rare because of changes linked to high intensity, continuous grazing that reduced the amount of light, fluffy fuel required for a fire. Presence of woody plants, little fine fuel and community fragmentation seem to be the causes of reduced fire frequency (Collins and Wallace, 1990). Grass biomass, fire frequency and fire size have decreased while the number and size of woody plants had increased. The woody plant seedlings found in grasslands are sensitive to fire (McPherson, 1995). Sensitivity is size dependent with mortality linked to small size. Many woody plants will not re-sprout if top killed and most are susceptible to fire mortality until they are fairly large. If encroaching woody plants do not produce seeds before the next fire, or do not re-sprout after being top killed by fire, these grasslands would remain relatively free of woody plants. Fire tolerant woody species would be suppressed by reoccurring fires and remain in the grassland, but at a small size (Archer, 1994). However, with a reduction of the grass fine fuel by heavy and constant herbivory, fire frequencies would decrease to zero, promoting establishment, growth and increased woody plant cover and density.

1.10. Herbivory Herbivores are really plant predators and damage plants by consuming parts. Herbivory can be aboveground or belowground or both, and the amount of damage is determined by the timing, amount of tissue eaten, and how often the plants are attacked (Crawley, 1997; Begon et al., 2006). As a result, herbivores can change interactions by altering a plant’s ability to obtain resources or by selectively eliminating a plant as a competitor. Herbivory can increase the number of gaps in a community, reduce aboveground and belowground biomass, and modify resource availability (Bush and Van Auken, 1995; Van Auken and Bush, 1997). Removal of biomass, may lead to changes in abundance and distribution through alteration in fecundity, and lack of re-growth or mortality (Harper, 1977; Crawley, 1997; Begon et al., 2006). Plant re-growth is usually reduced after one or more encounters with herbivores, and aboveground removal is usually associated with reductions in belowground growth and biomass (Detling et al., 1979; Van Auken and Bush, 1989). There are reports that this is not always the case in some systems (McNaughton et al., 1998) and aboveground re-growth, compensation or overcompensation after herbivory has been a contentious issue (McNaughton et al., 1998; Belsky et al., 1999; Begon et al., 2006). In the American southwest, reduction of grass density and cover by herbivory and the concomitant encroachment of woody plants, have coincided with or been preceded by intensification of grazing by large numbers of domestic herbivores (Bahre, 1991; Archer, 1994; Archer et al., 1995; Bahre, 1995). Alterations in the species composition of grasslands and reductions in herbaceous plant basal area, density, aboveground and belowground biomass are known to accompany chronic high levels of livestock grazing (Heitschmidt and Stuth, 1991). Grass herbivory at low intensity and frequency may cause little change in a grassland community, but at high intensity and frequency, coupled to reduced fire frequency, grasslands will become shrublands or woodlands (Knapp et al., 2008a; Knapp et al., 2008b).

Critical to the above indicated changes are the species consumed. When grasses are consumed selectively, they will be at a disadvantage to other plant species present. If browsers consume the shrubs or woody plants, the woody plants will be at a disadvantage (Van Auken and Smeins, 2008). In systems where woody plants are browsed, these plants remain small and their density remains low or static, but with removal of the browsers, plant size and density increase. Herbivores in African grasslands and savannas (grazers and browsers) have been shown to exert a major influence on woody plant distribution and abundance (see Van Auken, 2000; Van Auken and Smeins, 2008). In North America grasslands, defoliation of woody seedlings by rodents, lagomorphs and insects is an important source of mortality, particularly for P. glandulosa. Cynomys ludovicianus (black-tailed prairie dog) has been shown to consume seeds, pods and seedlings of P. glandulosa and to maintain their colony surface clear of seedlings and saplings (Weltzin et al., 1998). The unfortunate large-scale eradication of prairie dogs and reduction of colonies by 98% in the early 1900s may have removed an important constraint to woody plant establishment over a large area of the American southwest. However, the timing of this eradication was apparently after extensive encroachment of P. glandulosa into many semiarid grasslands. 1.11. Seed dispersal The presence of an inadequate number of seed dispersers during most of the Holocene is a theory presented to explain the maintenance of woody plant free grasslands during this time period. When this limitation was removed by the recent introduction of domestic livestock, woody plants spread into many types of grassland (Brown and Archer, 1989; Brown and Archer, 1999). However, a number of large and small native mammals are known to feed on the fruit of woody plants including P. glandulosa a native legume and they act as seed dispersal agents (Wilson, 1993; Kamp et al., 1998). P. glandulosa requires scarification, which may occur during mastication, and passage through the gut of cattle and various native species (Brown and Archer, 1989; Wilson, 1993; Kamp et al., 1998). The introduction of domestic herbivores may have increased the dispersal of P. glandulosa and other woody plant seeds, but many native herbivores did and still do the same thing. Prosopis and other woody plants were present in southwestern North American grasslands in about the same general areas where they were prior to increases in the number of domestic herbivores, except their local distribution, density and stature have increased with the increase in domestic herbivores. These increases in woody plant density did not require long distance seed dispersal by domestic herbivores and seeds could have been dispersed just as far by native species. 1.12. The consequences of encroachment The density, cover and biomass of certain woody species increased as the density, cover and biomass of many species of grasses and other herbaceous species decreased in the vast area of grassland where encroachment has been occurring (Knapp et al., 2008a; Knapp et al., 2008b). There was also a shift in allocation of plant biomass, carbon and nitrogen pools from largely below ground in grasslands to aboveground in developing woodlands (McKinley et al., 2008a). But what happened to the soil components and soil processes is not as clear. The distribution of soil nutrients seems to be less uniform and more variable compared to former grasslands (Schlessinger et al., 1990; Schlesinger et al., 1999; Throop and Archer, 2008). Higher concentrations of nutrients and organics are associated with the woody or shrub canopy rather that the

O.W. Van Auken / Journal of Environmental Management 90 (2009) 2931–2942

interspaces. High levels of soil nitrogen are associated with the canopy of encroaching woody legumes such as Prosopis and Acacia (Bush, 2008). However, where other woody species are encroaching, changes in soil resources are just being elucidated. Encroachment in xeric sites caused a decrease in aboveground net primary production, but in more mesic sites there was an increase in production because of high shrub leaf area (Knapp et al., 2008b). In new Juniperus woodlands, carbon and nitrogen stocks increased in plant biomass and soils with little change in availability of nitrogen (McKinley and Blair, 2008). Few changes in nitrogen cycle processes were detected in Juniperus woodlands recently converted from grasslands, including little change in microbiological activity (McKinley et al., 2008b). Soil temperatures were lower below the Juniperus canopies while soil moisture remained unchanged. These studies suggest the biggest changes are in the location of the biomass (belowground in the grassland versus aboveground in the woodland) and the form of the biomass (herbaceous versus woody). The herbaceous biomass would support the grazers while the woody biomass, depending on the species, would support some browsers, but many woody species would be difficult for the browsers to use. With the shift in plant biomass from herbaceous species to woody species, these former grassland communities can on longer support large populations of grazers and the pastoral economy as in the past (Campbell et al., 1997; Reynolds et al., 2007). There has been a shift in the management and economy in many areas of the American southwestern arid and semiarid grasslands from harvesting cattle and other grazing animals to harvesting white-tailed deer and other browsing ungulates (Doughty, 1983). However, there is some evidence that shrub or woody plant cover has started to stabilize, with little continued temporal expansion (Browning et al., 2008). Some additional evidence from northwestern Juniperus communities suggests fewer new woodland communities are establishing today compared to the recent past (Miller et al., 2005). This suggests that the period of greatest encroachment may be past but evidence is limited. 1.13. Specific mechanism to explain encroachment Seeds of woody plants have been introduced into North American grasslands since the earliest times. Seeds germinated, plants grew and established, but few matured and produced seed for the next generation. Something happened in many of these grasslands in the mid to late 1800s that allowed the woody plants to complete their life cycle and increase their local density and continue to spread. The specific mechanism involving the start of the rapid woody plant encroachment in the past 160 years seems to be the reduction of grass aboveground biomass. This would cause a reduction in the fire frequency because of the loss of fine fuel. Associated with this reduction would be high levels of surface soil disturbance. Without fires, reduction of aboveground grass biomass by chronic high levels of herbivory would lead to a reduction of belowground biomass because of reallocation of resources to leaves and stems. A cycle of loss grass biomass would follow including a reduction or decrease in the competitive ability of the grass (see (Van Auken, 2000)). Increased soil disturbance at the time grass cover was reduced would cause increased runoff and erosion leading to increases in heterogeneity of soil resources, especially water, nitrogen and probably phosphate (Schlessinger et al., 1990; Schlesinger et al., 1996; Schlesinger et al., 1999). Reduced surface soil resource levels, especially nitrogen, would favor establishment of many of the leguminous shrubs including members of the genus Acacia, Cercidium and Prosopis that have low soil nitrogen requirements. Other woody species that might be favored would include those with presumed low resource or

2939

nutrient requirements such as Larrea, Flourensia, Juniperus and some Quercus. As the woody plants establish and grow, surface soil resources would be partitioned differently, accumulating below the woody plant canopies. These clumps of woody plant are called resource islands, islands of fertility or islands of diversity and are where woody species and soil resources are concentrated in many of these communities (Bush and Van Auken, 1986; Archer et al., 1988; Schlessinger et al., 1990). Soil resources would be cycled in these islands of fertility making them favored sites for new species establishment and community succession. There would be a canopy or partial canopy in place, reducing light levels and soil surface temperatures and increasing the soil water content. New species of woody plants with different requirements could establish, further changing the composition and structure of these communities and increasing the difficulty for grass re-establishment and thus the return to an arid or semiarid grassland community. 1.14. Control methods In the past, woody plant or brush encroachment was treated as a problem associated with the woody species. The encroaching woody species were considered aliens and invaders with a high degree of aggressiveness. The anthropogenic terms suggested that these species had special powers, when in fact they were just growing extremely well in disturbances without competition. These are areas or conditions where these species normally grow very well. What was happening to the grass in the grassland or rangeland was not considered. To deal with what appeared to be a serious grassland and rangeland problem various mechanical (chaining, rolling, chopping or shredding) and chemical (various herbicide) treatments were used (Scifres, 1980). Other treatments were tried including fire and biological control without a lot of apparent success. Fire would not work with little fuel. In addition, woody plant mortality is size and temperature dependent, with large size plants being fire resistant. In addition, as soon as the grasses started to re-grow after treatment, usually after one or more rains, the animals (typically cattle) were put back on the treated area. The result of this type of treatment was an approximate 20 year cycle of treatment, grass re-growth, woody plant encroachment followed by re-treatment. More recently, combinations of treatments have been used with greater success but difficulties still occur and without fires, treatments are very expensive with limited success. Management usually includes some degree of mechanical or chemical treatment but success is still limited and reversing the process or going from a woodland or shrubland to grassland is complex and difficult (Ansley and Wiedemann, 2008). Biotic controls are usually more sophisticated, are usually combined with fire and may include genetic manipulation of some of the browsing species (Taylor, 2008). 2. Conclusions Interpreting recent past (160 years ago to present) plant community composition and changes with information that is little more than anecdotal is a perplexing task. Predicting future plant community structure based on limited understanding of current ecosystems and potential future climatic conditions is also difficult. In spite of these limitations, I will present what seems to be a reasonable interpretation of causes and consequences of woody plant encroachment. There is no doubt that the climate of the Earth is changing probably because the orbit of the Earth is becoming more circular and because of increasing levels of atmospheric CO2. These

2940

O.W. Van Auken / Journal of Environmental Management 90 (2009) 2931–2942

changes, both natural and anthropogenic, will result in very different distributions of plant communities and very different community structure and composition on the Earth surface. Changes will occur as the climate becomes more moderate. With higher temperatures, plant and animal species will migrate. Temperate species in the northern hemisphere will move north and up in elevation as cold tolerant species do the same. Tropical species will likely expand their ranges away from the equator into more temperate regions and also into higher elevations as the climate moderates. These current migrations began with the start of warming at the end of the most recent glacial period, the beginning of the Holocene, 11,000–12,500 years ago. However, the plant communities of the arid and semiarid grasslands of southwestern North America and grasslands in other parts of the world seem to be responding to additional factors or conditions. The rate of encroachment of native woody species into these grasslands appears to have increase about 160 years ago. This was before the current major increase in atmospheric CO2 and the concomitant anthropogenically caused rise in temperature. Reduced competition from native grasses (because of domestic herbivory), dispersal of seeds of woody plants by domestic animals, changes in invertebrate, rodent and lagomorph populations all seem to modify the current rate of community change, but are not the main cause of the encroachment. The start of the recent encroachment was associated with the introduction of millions of domestic animals into the grasslands of southwestern North America and high levels of constant herbivory. This domestic herbivory reduced the above ground grass biomass, the fine fuel, required for large-scale ecosystem fires. With the reduction of the fuel there was a concomitant decrease in grassland fires. These new ecological conditions favored the growth of the woody plants and not the grasses that had dominated these communities for thousands of years. The process continues today and will continue into the future until woody plans have encroached into all or most of the grasslands or until the relationship between fires, woody plants and grassland communities is accepted. The encroachment of woody species follows when herbivores remove light-fluffy fuel in the form of grass biomass and the grasslands cannot burn because there is little fine fuel remaining to burn. The fires would have killed or top killed the woody plants and prevented an increase in woody plant density in these grassland communities. However, grassland fires certainly interact with herbivory, rainfall and temperature to complicate the interpretation of woody plant encroachment into arid and semiarid grassland ecosystems. Can the process of encroachment be slowed or reversed? This will be an arduous task. These dryland communities in southwestern North America and in other places have changed over the past 160 years. In some areas of the world the process started much sooner and the time span has been longer. The communities cannot be changed back from savanna or woodland to grasslands overnight. Making the changes or reversing the current encroached ecosystems and ecosystem processes will be difficult and expensive. It will also require a change in mindset of those who are harvesting the grass biomass through grazing in these marginal ecosystems. Sustainability of these ecosystems should be the overall goal. Some grazing can be sustained in these dryland ecosystems, but these communities cannot sustain a high level of constant grass biomass removal and the grasslands cannot be sustained without fire. Acknowledgements I appreciate the help with some of the photographs, discussion of the ideas included and insight into some of the problems in

biological communities caused by global change phenomena offered by D.C. McKinley. I also appreciate the help by J.K. Bush when I was dealing with the complexities of this manuscript. References Allen, C., Betancourt, J., Swetnam, T., 2002. Range expansion of woody plants on the Colorado Plateau. In: Grahame, J.D., Sisk, T.D. (Eds.). http://www.cpluhna. nau.edu/. Ansley, R.J., Wiedemann, H.T., 2008. Reversing the woodland steady state: vegetation responses during restoration of Juniperus-dominated grasslands with chaining and fire. In: Van Auken, O.W. (Ed.), Western North American Juniperus Communities: A Dynamic Vegetation Type. Springer, New York, pp. 272–290. Archer, S., 1994. Woody plant encroachment into southwestern grasslands and savannas: rates, patterns and proximate causes. In: Vavra, M., Laycock, W.A., Pieper, R.D. (Eds.), Ecological Implications of Livestock Herbivory in the West. Society for Range Management, Denver, pp. 13–69. Archer, S., Scifres, C., Bassham, C.R., Maggio, R., 1988. Autogenic succession in a subtropical savanna: conversion of grassland to thorn woodland. Ecol. Monogr. 52, 111–127. Archer, S.R., Schimel, D.S., Holland, E.H., 1995. Mechanisms of shrubland expansion: land use, climate or CO2. Clim. Change 29, 91–99. Bahre, C.J., 1991. A Legacy of Change: Historic Human Impact on Vegetation on the Arizona Borderlands. University of Arizona Press, Tucson. Bahre, C.J., 1995. Human impacts on the grasslands of southeastern Arizona. In: McClaran, M.P., Van Devender, T.R. (Eds.), The Desert Grassland. The University of Arizona Press, Tucson, pp. 230–264. Bahre, C.J., Shelton, M.L., 1993. Historic vegetation change, mesquite increases, and climate in southeastern Arizona. J. Biogeogr. 20, 489–504. Bailey, R.G., 1978. Description of the Ecoregions of the United States. USDA, Forest Service, Intermountain Region, Ogden, UT. Bartolome, J.W., 1993. Application of herbivory optimization theory to rangelands of the western United States. Ecol. Appl. 3, 27–39. Begon, M., Townsend, C.R., Harper, J.L., 2006. Ecology: From Individuals to Ecosystems. Blackwell Publishing, Malden, MA. Belsky, A.J., 1986. Does herbivory benefit plants? A review of the evidence. Am. Nat. 127, 870–892. Belsky, A.J., 1996. Viewpoint: western juniper expansion: is it a threat to arid northwestern ecosystems? J. Range Manag. 49, 53–59. Belsky, A.J., Blumenthal, D.M., 1997. Effects of livestock grazing on stand dynamics and soils in upland forests of the interior west. Conserv. Biol. 11, 315–327. Belsky, A.J., Matzke, A., Uselman, S., 1999. Survey of livestock influences on stream and riparian ecosystems in the western United States. J. Soil Water Conserv. 54, 419–431. Berger, A., Loutre, M.F., 2002. An exceptionally long interglacial ahead? Science 297, 1287–1288. Berner, R.A., 2005. The rise of trees and how they changed Paleozoic atmospheric CO2, climate, and geology. In: Ehleringer, J.R., Cerling, T.E., Dearing, M.D. (Eds.), A History of Atmospheric CO2 and Its Effects on Plants, Animals, and Ecosystems. Springer, New York, pp. 1–7. Betancourt, J.L., Van Devender, T.R., Martin, P.S., 1990. Synthesis and prospectus. In: Betancourt, J.L., Van Devender, T.R., Martin, P.S. (Eds.), Packrat Middens: The last 40,000 years of Biotic Change. University of Arizona Press, Tucson, pp. 435–447. Breshears, D.D., 2008. Structure and function of woodland mosaics: consequences of patch-scale heterogeneity and connectivity along the grassland-forest continuum. In: Van Auken, O.W. (Ed.), Western North American Juniperus Communities: A Dynamic Vegetation Type. Springer, New York, pp. 58–92. Breshears, D.D., Cobb, N.S., Rich, P.M., Price, K.P., Allen, C.D., 2005. Regional vegetation die-off in response to global-change type drought. Proc. Nat. Acad. Sci. U.S.A. 102, 15144–15148. Briggs, J.M., Knapp, A.K., Blair, J.M., Heisler, J.L., Hoch, G.A., Lett, M.S., McCarron, K., 2005. An ecosystem in transition: woody plant expansion into mesic grassland. Bioscience 55, 243–254. Briske, D., 1993. Grazing optimization: a plea for a balanced perspective. Ecol. Appl. 3, 24–26. Brock, M., Wade, M., Pysek, P., Green, D., 1997. Plant Invasions: Studies from North America and Europe. Backhuys, Leiden. Brown, J.R., Archer, S., 1989. Woody plant invasion of grasslands: establishment of honey mesquite (Prosopis glandulosa var. glandulosa) on sites differing in herbaceous biomass and grazing history. Oecologia 80, 19–26. Brown, J.R., Archer, S., 1999. Shrub invasion of grassland: recruitment is continuous and not regulated by herbaceous biomass or density. Ecology 80, 2385–2396. Browning, D.M., Archer, S.R., Asner, G.P., McClaran, M.P., Wessman, C.A., 2008. Woody plants in grasslands: post-encroachment stand dynamics. Ecol. Appl. 18, 928–944. Buffington, L.C., Herbel, C.H., 1965. Vegetational changes on a semidesert grassland range from 1858 to 1963. Ecol. Monogr. 35, 139–164. Bush, J.K., 2008. Soil nitrogen and carbon after twenty years of riparian forest development. Soil Sci. Soc. Am. J. 72, 815–822. Bush, J.K., Van Auken, O.W., 1986. Changes in nitrogen, carbon, and other surface soil properties during secondary succession. Soil Sci. Soc. Am. J. 50, 1597–1601. Bush, J.K., Van Auken, O.W., 1995. Woody plant growth related to planting time and clipping of a C4 grass. Ecology 76, 1603–1699.

O.W. Van Auken / Journal of Environmental Management 90 (2009) 2931–2942 Cabral, A.C., De Miguel, J.M., Rescia, A.J., Schmitz, M.F., Pineda, F.D., 2003. Schrub encroachment in Argentinean savannas. J. Veg. Sci. 14, 145–152. Campbell, B.D., Stafford-Smith, D.M., McKeon, G.M., 1997. Elevated CO2 and water supply interactions in grasslands: a pastures and rangelands management perspective. Global Change Biol. 3, 177–187. Cole, M.M., 1986. The Savannas: Biogeography and Geo Botany. Academic Press, New York. Collins, S.L., Wallace, L.L., 1990. Fire in North American Tallgrass Prairies. University of Oklahoma Press, Norman, Oklahoma. Correll, D.S., Johnston, M.C., 1970. Manual of the Vascular Plants of Texas. Texas Research Foundation, Renner, Texas. Crawley, M.J., 1997. Plant–herbivore dynamics. In: Crawley, M. (Ed.), Plant Ecol. Blackwell Science Ltd., Oxford, pp. 401–474. Delcourt, P.A., Delcourt, H.R., Webb, T., 1983. Dynamic plant ecology: the spectrum of vegetation change in space and time. Quat. Sci. Rev. 1, 153–175. Detling, J.K., Dyer, M.I., Winn, D.T., 1979. Net photosynthesis, root respiration, and regrowth of Bouteloua gracilis following simulated grazing. Oecologia 41, 127–134. Dick-Peddie, W.A., 1993. New Mexico Vegetation: Past, Present and Future. University of New Mexico Press, Albuquerque. Doughty, R.W., 1983. Wildlife and Man in Texas: Environmental Change and Conservation. Texas A & M University Press, College Station, Texas. Dyer, M.I., Turner, C.L., Seastedt, T.R., 1993. Herbivory and its consequences. Ecol. Appl. 3, 10–16. Ehleringer, J.R., 2005. The influence of atmospheric CO2, temperature, and water on the abundance of C3/C4 taxa. In: Ehleringer, J.R., Cerling, T.E., Dearing, M.D. (Eds.), A History of Atmospheric CO2 and Its Effects on Plants, Animals, and Ecosystems. Springer, New York, pp. 214–231. Enserink, M., 1999. Biological invaders sweep in. Science 285, 1834–1836. Fensham, R.J., Fairfax, R.J., Archer, S.R., 2005. Rainfall, land use and woody vegetation cover change in semi-arid Australian savanna. J. Ecol. 93, 596–606. Fuhlendorf, S.D., Archer, S.R., Smeins, F.E., Engle, D.M., Taylor Jr., C.A., 2008. The combined influence of grazing, fire, and herbaceous productivity on tree–grass interactions. In: Van Auken, O.W. (Ed.), Western North American Juniperus Communities: A Dynamic Vegetation Type. Springer, New York, pp. 219–238. Gardener, C.J., McIvor, J.G., Williams, J., 1990. Dry tropical rangelands: solving one problem and creating another. Proc. Ecol. Soc. Aust. 16, 279–286. Geist, H.J., Lambin, E.F., 2004. Dynamic causal patterns of desertification. Bioscience 54, 817. GLP, 2005. Global land project-science plan and implementation strategy. International Geosphere Biosphere Program report No. 53/International Human Dimensions Programme Report No. 19, Stockholm. Grover, H.D., Musick, H.B., 1990. Shrubland encroachment in southern New Mexico, USA: an analysis of desertification processes in the American southwest. Clim. Change 17, 305–330. Harper, J.L., 1977. Population Biology of Plants. Academic Press, New York. Hastings, J.R., Turner, R.M., 1965. The Changing Mile. University of Arizona Press, Tucson. Havstad, K.M., Schlesinger, W.H., Huenneke, L.F., 2006. Structure and Function of a Chihuahuan Desert Ecosystem: The Jornada Basin LTER. Oxford University Press, New York. Heitschmidt, R.K., Stuth, J.W., 1991. Grazing Management: An Ecological Perspective. Timberline Press, Portland. Hoch, G.A., Briggs, J.M., Johnson, L.C., 2002. Assessing the rate, mechanism and consequences of tallgrass prairie to Juniperus virginiana forest. Ecosystems 6, 578–586. House, J., Archer, S., Breshears, D.D., Scholes, R.J., Nceas, T.G.I.P., 2003. Conundrums in mixed woody-herbaceous plant systems. J. Biogeogr. 30, 1763–1777. Humphrey, R.R., 1958. The desert grassland: a history of vegetational change and an analysis of causes. Bot. Rev. 24, 193–252. Idso, S.B., 1992. Shrubland expansion in the American southwest. Clim. Change 22, 85–86. Imbrie, J., Imbrie, J.Z., 1980. Modeling the climatic response to orbital variations. Science 207, 943–953. Imbrie, J., Imbrie, K.P., 1979. Ice Ages: Solving the Mystery. Enslow, Short Hills, New Jersey. Inglis, J.M., 1964. A History of Vegetation on the Rio Grande Plain. Texas Parks and Wildlife Department, Austin, Texas. Johnson, H.B., Polley, H.W., Mayeux, H.S., 1993. Increasing CO2 and plant–plant interactions: effects on natural vegetation. Vegetation 104-105, 157–170. Kaiser, J., 1999. Stemming the tide of invading species. Science 285, 1836–1841. Kamp, B.A., Ansley, R.J., Tunnell, T.R., 1998. Survival of mesquite seedlings emerging from cattle and wildlife feces in a semi-arid grassland. Southwest. Nat. 43, 300–312. Knapp, A.K., Briggs, J.M., Collins, S.L., Archer, S.R., 2008b. Shrub encroachment in North American grasslands: shifts in growth form dominance rapidly alters control of ecosystem carbon inputs. Global Change Biol. 14, 615–623. Knapp, A.K., McCarron, J.K., Silletti, G.A., Hoch, G.l., Heisler, M.S., Lett, J.M., Blair, J.M., Briggs, J.M., Smith, M.D., 2008a. Ecological consequences of the replacement of native grassland by Juniperus virginiana and other woody plants. In: Van Auken, O.W. (Ed.), Western North American Juniperus Communities: A Dynamic Vegetation Type. Springer, New York, pp. 156–169. Lauenroth, W.K., 1979. Grassland primary production. In: French, N.R. (Ed.), Perspectives in Grassland Ecology: Results and Applications of the US/IBP Grassland Biome Study. Springer-Verlag, New York, pp. 3–24.

2941

Long, S.P., 1991. Modification of the response of photosynthetic productivity to rising temperature by atmospheric CO2 concentrations: has its importance been underestimated? Opinion. Plant Cell Environ. 14, 729–740. Louda, S.M., Keeler, K.H., Holt, R.D., 1990. Herbivory influences on plant performance and competitive interactions. In: Grace, J.B., Tilman, D. (Eds.), Perspectives on Plant Competition. Academic Press, New York, pp. 413–444. MacDonald, I.A.W., 1989. Man’s role in changing the face of southern Africa. In: Huntley, B.J. (Ed.), Biotic Diversity in Southern Africa: Concepts and Conversions. Oxford University Press, Cape Town, pp. 51–78. Mackenzie, F.T., 2003. Our Changing Planet: an Introduction to Earth System Science and Global Environmental Change. Prentice Hall-Pearson Education, Inc., Upper Saddle River, New Jersey. Malakoff, D., 1999. Fighting fire with fire. Science 285, 1841–1843. Martin, P.S., 1999. Deep history and a wilder West. In: Robichaux, R.H. (Ed.), Ecology of the Sonoran Desert plants and Plant Communities. The University of Arizona Press, Tucson, pp. 255–290. Mayeux, H.S., Johnson, H.B., Polley, H.W., 1991. Global change and vegetation dynamics. In: James, L.F., Evans, J.O., Ralphs, M.H., Sigler, B.J. (Eds.), Noxious Range Weeds. Westview Press, Boulder, Colorado, pp. 62–74. McClaran, M.P., Van Devender, T.R., 1995. The Desert Grassland. University of Arizona Press, Tucson. McDowell, P.F., Webb III, T., Bartlein, P.J., 1995. Long-term environmental change. In: Powell, T.M., Steele, J.H. (Eds.), Ecological Time Series. Chapman & Hall, New York, pp. 327–370. McKinley, D.C., Blair, J.M., 2008. Woody plant encroachment by Juniperus virginiana in a mesic native grassland promotes rapid carbon and nitrogen accrual. Ecosystems 11, 454–468. McKinley, D.C., Morris, M.D., Blair, J.M., Johnson, L.C., 2008a. Altered ecosystem processes as a consequence of Juniperus virginiana L. encroachment into North American tallgrass prairie. In: Van Auken, O.W. (Ed.), Western North American Juniperus Communities: A Dynamic Vegetation Type. Springer, New York, pp. 170–187. McKinley, D.C., Rice, C.W., Blair, J.M., 2008b. Conversion of grassland to coniferous woodland has limited effects on soil nitrogen cycle processes. Soil Biol. Biochem. 40, 2627–2633. McNaughton, S.J., 1993. Grasses and grazers, science and management. Ecol. Appl. 3, 17–20. McNaughton, S.J., Banyikwa, F.F., McNaughton, M.M., 1998. Root biomass and productivity in a grazing system: the Serengeti. Ecology 79, 587–592. McPherson, G.R., 1995. The role of fire in the desert grassland. In: McClaran, M.P., Van Devender, T.R. (Eds.), The Desert Grassland. The University of Arizona Press, Tucson, pp. 131–151. McPherson, G.R., 1997. Ecology and Management of North American Savannas. University of Arizona Press, Tucson. McPherson, G.R., Wright, H.W., Wester, D.B., 1988. Patterns of shrub invasion in semiarid Texas grasslands. Am. Midl. Nat. 120, 391–397. MEA, 2005. Millennium Ecosystem Assessment-Ecosystems and Human Wellbeing: Desertification Synthesis. World Resources Institute, Washington, D.C. Milankovitch, M., 1998. Canon of Insolation and the Ice Age Problem. Alven Global. Miller, R.F., Bates, J.D., Svejcar, T.J., Pierson, F.B., Eddleman, L.E., 2005. Biology, Ecology and Management of Western Juniper (Juniperus occidentalis). Technical bulletin 152. Agriculture Experiment Station, Oregon State University, Corvalis, Oregon. Miller, R.F., Tausch, R.J., McArthur, E.D., Johnson, D.D., Sanderson, S.C., 2008. Age Structure and Expansion of Pinion–juniper Woodlands: a Regional Perspective in the Intermountain West. Research Paper RMRS-RP-69. USDA, Forrest Service, Rocky Mountain Research Station, Fort Collins, Colorado, pp. 1–15. Miller, R.F., Wigand, P.E., 1994. Holocene changes in semiarid pinyon–juniper woodlands: response to climate, fire, and human activities in the Great Basin. Bioscience 44, 465–474. Misra, R., 1983. Indian savannas. In: Bourliere, F. (Ed.), Tropical Savannas. Elsevier, New York, pp. 151–166. Mlambo, D., Nyathi, P., Mapaure, I., 2005. Influence of Colophospermum mopane on surface soil properties and understory vegetation in a southern African savanna. For. Ecol. Manag. 212, 394–404. Mooney, H.A., Drake, B.G., Luxmoore, R.J., Oechel, W.C., Pitelka, L.F., 1991. Predicting ecosystem response to elevated CO2 concentrations. Bioscience 41, 96–104. Mooney, H.A., Drake, J.A., 1986. Ecology of Biological Invasions of North America and Hawaii. Springer-Verlag, New York. Morgan, J.A., Milchunas, D.G., LeCain, D.R., West, M., Mosier, A.R., 2007. Carbon dioxide enrichment alters plant community structure and accelerates shrub growth in the shortgrass steppe. Proc. Nat. Acad. Sci. U.S.A. 104, 14724–14729. Neff, J.C., Ballantyne, A.P., Farmer, G.L., Mahowald, N.M., Conroy, J.L., Landry, C.C., Overpeck, J.T., Painter, T.H., Lawrence, C.R., Reynolds, R.L., 2008. Increasing eolian dust deposition in the western United States linked to human activity. Nature Geosci. 1, 189–195. Noy-Meir, I., 1993. Compensating growth of grazed plants and its relevance to the use of rangelands. Ecol. Appl. 3, 32–34. Painter, E.L., Belsky, A.J., 1993. Application of herbivore optimization theory to rangelands of the western United States. Ecol. Appl. 3, 2–9. Patten, D.T., 1993. Herbivore optimization and overcompensation: does native herbivory of western rangelands support these theories? Ecol. Appl. 3, 35–36. Perry, R.A., 1970. Arid shrublands and grasslands. In: Moore, R.M. (Ed.), Australian Grasslands. Australian National University Press, Canberra, pp. 246–259.

2942

O.W. Van Auken / Journal of Environmental Management 90 (2009) 2931–2942

Peters, D., Bestelmeyer, P.C., Brandon, T., Herrick, J.E., Fredrickson, E.L., Monger, H.C., Havstad, K.M., 2006. Disentangling complex landscapes: new insights into arid and semiarid system dynamics. Bioscience 56, 491–501. Petit, J.R., Jouzel, J., Raynaud, D., Barkof, N.L., Barnola, J.M., Basile, I., Bender, M., Chappellaz, J., Davis, M., Delaygue, G., Delmotte, M., Kotlyakov, V.M., Legrand, M., Lipenkov, V.Y., Lorius, C., Pepin, L., Ritz, C., Saltzman, E., Stievenard, M., 1999. Climate and atmospheric history of the past 420,000 years from the Vostok ice core, Antarctica. Nature 399, 429–436. Pimm, S.L., Russell, G.J., Gittleman, J.L., Brooks, T.M., 1995. The future of biodiversity. Science 269, 347–350. Polley, H.W., Johnson, H.B., Mayeux, H.S., 1992. Carbon dioxide and water fluxes of C3 annuals and C3 and C4 perennials at subambient CO2 concentrations. Func. Ecol. 6, 693–703. Poorter, H.M., Navas, 2003. Plant growth and competition at elevated CO2: on winners, loosers and functional groups. New Phytol. 157, 175–198. Reynolds, J.F., Smith, D.M.S., Lambin, E.F., Turner II, B.L., 2007. Global desertification: building a science for dryland development. Science 316, 847–851. Roques, K.G., O’Connor, T.G., Watkinson, A.R., 2001. Dynamics of shrub encroachment in an African savanna: relative influences of fire, herbivory, rainfall and density dependence. J. Appl. Ecol. 38, 268–280. Ryniker, K.A., Bush, J.K., Van Auken, O.W., 2006. Structure of Quercus gambelii communities in the Lincoln National Forest, New Mexico, USA. For. Ecol. Manag. 233, 69–77. Schlesinger, W.H., Abrahams, A.D., Parsons, A.J., Wainwright, J., 1999. Nutrient losses in runoff from grassland and shrubland habitats in southern New Mexico: rainfall simulation experiments. Biogeochemistry 45, 21–34. Schlesinger, W.H., Raikes, J.A., Hartley, A.E., Cross, A.F., 1996. On the spatial pattern of soil nutrients in desert ecosystems. Ecology 77, 364–374. Schlessinger, W.H., Reynolds, J.F., Cunningham, G.L., Huenneke, L.F., Jarrell, W.H., Virginia, R.A., Whitford, W.G., 1990. Biological feedbacks in global desertification. Science 247, 1043–1048. Schmutz, E.M., Smith, E.L., Ogden, P.R., 1991. Desert grassland. In: Coupland, R.T. (Ed.), Natural Grasslands: Introduction and Western Hemisphere. Elsevier, Amsterdam, pp. 337–362. Scholes, R.J., Walker, B.H., 1993. An African Savanna: Synthesis of the Nylsvley Study. Cambridge University Press, Cambridge. Scifres, C.J., 1980. Brush Management Principles and Practices for Texas and the Southwest. Texas A&M University Press, College Station. Seager, R., Ting, M., Held, I., Kushnir, Y., Lu, J., Vecchi, G., Huang, H.-P., Harnik, N., Leetmaa, A., Lau, N.-C., Li, C., Velez, J., Naik, N., 2007. Model projections of an imminent transition to a more arid climate in southwestern North America. Science 316, 1181–1184. Sharma, R., Dakshini, K.M.M., 1991. A comparative assessment of the ecological effects of Prosopis cineraria and Prosopis juliflora on the soil of revegetated spaces. Vegetation 96, 87–96. Shaw, M.R., Huxman, T.E., Lund, C.P., 2005. Modern and future semi-arid and arid ecosystems. In: Ehleringer, J.R., Cerling, T.E., Dearing, M.D. (Eds.), A History of Atmospheric CO2 and Its Effects on Plants, Animals and Ecosystems. Springer, New York, pp. 415–440. Silva, J.F., Zambrano, A., Farinas, M.R., 2001. Increase in the woody component of seasonal savannas under different fire regimes in Calabozo, Venezuela. J. Biogeogr. 28, 977–983. Simmons, M.T., Archer, S.R., Ansley, R.J., Teague, W.R., 2007. Grass effects on tree (Prosopis glandulosa) growth in a temperate savanna. J. Arid Env. 69, 212–227.

Solomon, S., Plattner, G.-K., Knutti, R., Friedlingstein, P., 2009. Irreversible climate change due to carbon dioxide emissions. Proc. Nat. Acad. Sci. U.S.A. 106, 1704– 1709. Soriano, A., 1979. Distribution of grasses and grasslands in South America. In: Numata, M. (Ed.), Ecology of Grasslands and Bamboolands in the World. Dr. W. Junk, Boston, pp. 84–91. Stromberg, J.C., Gengarelly, L., Rogers, B.F., 1997. Exotic herbaceous species in Arizona’s riparian ecosystems. In: Brock, M., Wade, M., Pysek, P., Green, D. (Eds.), Plant Invasions: Studies from North America and Europe. Backhuys, Leiden, pp. 45–57. Swetnam, T.W., Betancourt, J.L., 1990. Fire-southern oscillation relations in the southwestern United States. Science 249, 1017–1020. Taylor Jr., C.A., 2008. Ecological consequences of using prescribed fire and herbivory to manage Juniperus encroachment. In: Van Auken, O.W. (Ed.), Western North American Juniperus Communities: A Dynamic Vegetation Type. Springer, New York, pp. 239–252. Throop, H.L., Archer, S.R., 2008. Shrub (Prosopis velutina) encroachment in a semidesert grassland: spatial–temporal changes in soil organic carbon and nitrogen pools. Global Change Biol. 14, 1–12. Van Auken, O.W., 2000. Shrub invasions of North American semiarid grasslands. Ann. Rev. Ecol. Syst. 31, 197–215. Van Auken, O.W., Bush, J.K., 1989. Prosopis glandulosa growth: influence of nutrients and simulated grazing of Bouteloua curtipendula. Ecology 70, 512–516. Van Auken, O.W., Bush, J.K., 1997. Growth of Prosopis glandulosa in response to changes in aboveground and belowground interference. Ecology 78, 1222– 1229. Van Auken, O.W., Smeins, F., 2008. Western North American Juniperus communities: patterns and causes of distribution and abundance. In: Van Auken, O.W. (Ed.), Western North American Juniperus Communities: A Dynamic Vegetation Type. Springer, New York, pp. 3–18. Van Devender, T.R., 1995. Desert grassland history: changing climates, evolution, biography, and community dynamics. In: McClaran, M.P., Van Devender, T.R. (Eds.), The Desert Grassland. University of Arizona Press, Tucson, pp. 68–99. van Klinken, R.D., Campbell, S.D., 2001. The biology of Australian weeds: Prosopis L. species. Plant Prot. Quart. 16, 2–20. Walker, H., 1971. Ecology of Tropical and Subtropical Vegetation. Oliver and Boyd, Edinburgh, UK. Wayne, E.R., Van Auken, O.W., 2008. Comparison of the understory vegetation of Juniperus woodlands. In: Van Auken, O.W. (Ed.), Western North American Juniperus Communities: A Dynamic Vegetation Type. Springer, New York, pp. 93–110. Wells, P.V., 1965. Scarp woodlands, transported grassland soil, and concepts of grassland climate in the Great Plains. Science 148, 246–249. Weltzin, J.F., Archer, S., Heitschmidt, R.K., 1998. Defoliation and woody plant (Prosopis glandulosa) seedling regeneration: potential vs realized herbivory tolerance. Plant Ecol. 138, 127–135. Weltzin, J.F., McPherson, G.R., 1999. Facilitation of conspecific seedling recruitment and shifts in temperate savanna ecotones. Ecol. Monogr. 69, 513–534. Wessman, C., Archer, S., Johnson, L., Asner, G., 2004. Woodland expansion in US grasslands: assessing land-cover change and biogeochemical impacts. In: Gutman, G., Janetos, A.C., Justice, C.O. (Eds.), Land Change Science: Observing, Monitoring and Understanding Trajectories of Change on the Earth’s Surface. Kluwer Academic Publishers, Dordrecht, pp. 185–208. Wilson, M., 1993. Mammals as seed-dispersal mutualists in North America. Oikos 67, 159–176.