Soil moisture correlates with shrub–grass association in the Chihuahuan Desert

Catena 107 (2013) 71–79

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Soil moisture correlates with shrub–grass association in the Chihuahuan Desert Giora J. Kidron a,⁎, Vincent P. Gutschick b a b

Institute of Earth Sciences, The Hebrew University, Givat Ram Campus, Jerusalem 91904, Israel Dept. of Biology, New Mexico State University, P.O. Box 30003, Las Cruces, NM 88003-8003, USA

a r t i c l e

i n f o

Article history: Received 22 July 2012 Received in revised form 31 January 2013 Accepted 4 February 2013 Keywords: Competition Facilitation Evaporation Biological soil crust

a b s t r a c t The current ongoing expansion of shrubs at the expense of grasses in the Chihuahuan Desert is largely believed to result from a shift in competition between shrubs and grasses. Both groups, shrubs and grasses, occupy a gently sloping silty (loessial) soil within the Jornada Experimental Range (JER), northern Chihuahuan Desert, NM, USA. At the sites dominated by tarbush (Flourensia cernua), the shrubs seem to be randomly distributed in between barren surfaces, surfaces covered by biocrusts, and dense stands of burrograss (Scleropogon brevifolius), which are mainly confined to 2–4 m diameter and 0.02–0.1 m deep depressions. Often, the shrubs are accompanied by a dense stand of tobosa grass (Pleuraphis mutica) that occupies the under-canopy. In order to study shrub– grass interactions, four habitats were defined: playa habitats with bare silty material (PL), habitats with mature biocrust (CR), habitats with burrograss mainly confined to depressions (DP) and tarbush shrubs with a dense population of tobosa grass under the shrub canopy (UC). Periodic moisture measurements at the 0–30 cm soil and potential evaporation took place during the summers of 1999 and 2009. In addition, the grass biomass in these habitats was monitored. Moisture content and grass biomass followed the pattern PL b CRb DPb UC, while the evaporation rate followed the pattern PL ≈CR> DP> UC. The findings point to a close link between grass biomass and the soil moisture content, with the high grass biomass at DP explained by runoff contribution, while that of UC by the higher moisture content mainly attributed to the reduced evaporation under shading conditions. The co-existence of shrubs and under-canopy grass does not support the notion of high competition between both groups of plants. It may rather indicate facilitation. The close link between soil moisture content and the grass habitats also suggests that lower water availability during summertime may explain current grass scarcity in JER. © 2013 Elsevier B.V. All rights reserved.

1. Introduction The ongoing expansion of C3 winter- (and spring) adapted shrubs at the expense of summer-adapted C4 grasses in the Chihuahuan Desert has attracted the attention of the scientific community. Grass predominated during the 19th century and the open areas in New Mexico were defined as desert plains grasslands (Clements, 1934). Ever since, grass cover has decreased dramatically. According to Buffington and Herbel (1965), while “good grass” was present in >90% of the Jornada (northern Chihuahuan Desert, New Mexico, USA) in 1858, it decreased to b 25% in 1963. According to Gibbens et al. (2005) grass declined to b7% in the Jornada in the last 150 years. Whereas the exact grass cover at the end of the 19th century cannot be exactly assessed, mainly due to the different methods employed (Buffington

⁎ Corresponding author. Tel.: +972 2 676 7271. E-mail address: [email protected] (G.J. Kidron). 0341-8162/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.catena.2013.02.001

and Herbel, 1965; Peters and Gibbens, 2006), all researchers agree that the shift in vegetation was of major magnitude (Peters et al., 2006). Some researchers claim that severe droughts are responsible for the vegetation shift (Bestelmeyer et al., 2006; Lohmiller, 1963). Yet, no persistent long-term decrease in precipitation has been observed (Conley et al., 1992; Friedman, 1957; Paulsen, 1956; Wainwright, 2006). Other researchers have attributed the shift to overgrazing that impacts the competitive capability of the grass (Campbell, 1929; Jardine and Forsling, 1922; Schlesinger et al., 1990; Scholes and Archer, 1997), to the reduction in fire frequency (as a result of reduced fuel load by grazing) which suppresses grass establishment (Harrington and Hodgkinson, 1986; Peters et al., 2006) or to CO2 increase that renders an advantage to C3 shrubs by altering photosynthetic rates and wateruse efficiency (Cole and Monger, 1994). Buffington and Herbel (1965), Phillips and MacMahon (1981) and later Schlesinger et al. (1990) and Jurena and Archer (2003) proposed that shrub competition with grass may explain shrub encroachment into the northern Chihuahuan Desert. According to these authors, competition is intensified due to overgrazing. This hypothesis has taken the forefront in attempts to explain the phenomenon. By

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creating stress on palatable grasses, grazing facilitates the establishment of shrubs, while creating grass-barren areas inbetween the shrubs. Runoff generated at these barren areas is responsible for nutrient transfer from the intershrub to the shrub habitats. By concentrating nutrients and above all nitrogen under the canopy (Cross and Schlesinger, 1999; Kieft et al., 1998; Schlesinger et al., 1996), a positive feedback mechanism is created as the shrubs deplete the intershrub areas from essential nutrients (Abrahams et al., 2006) necessary in turn for successful grass re-establishment (Schlesinger et al., 1990). Nevertheless, although considered the leading hypothesis accounting for vegetation shift in the northern Chihuahuan Desert and in other deserts (Grover and Musick, 1990; Peters et al., 2006), no direct proof has yet to be provided for the long-lasting effect of overgrazing and the hazardous effect of shrub–grass competition on grass depletion. Moreover, convincing evidence has yet to be provided regarding the cardinal effect of nutrients. Indeed, following water, nutrients and above all nitrogen are considered the second limiting factor in deserts (Noy-Meir, 1973). However, the nutrient effect is principally linked to plant productivity, as reflected by its biomass. While water explains over 80–90% of plant productivity (Sala et al., 1988; Sims and Singh, 1978; Sims et al., 1978), additional nutrients increase productivity (Gough et al., 2000). Yet, while all agree regarding the role played by water (and nutrients) in increasing the plant biomass (Noy-Meir, 1973), and while numerous works highlighted the essential role of water in seedling emergence (Beatley, 1969; Grover and Musick, 1990), convincing data has yet to be published regarding seedling impediment due to nutrient shortage. Contrarily, numerous experiments conducted in US deserts concluded that conditional upon adequate temperatures, seedling emergence is solely linked to water availability (Beatley, 1969; Brown and Archer, 1989; Grover and Musick, 1990; Went, 1949, 1973). Furthermore, no indication of shortage in N in leaves of desert plants was found (Schlesinger et al., 2006) and no increase in seedling emergence was recorded on or near dung (Brown and Archer, 1989). Moreover, nutrient examination at habitats with different grass species failed to detect significant differences (Devine et al., 1998), while the removal of either shrubs or grasses had a negligible effect (if any) on the other group (Huenneke and Noble, 1996; Huenneke and Schlesinger, 2006; Villagra and Felkert, 1997). These findings point out that aridification and shrub–grass distribution due to a shortage in nutrients are highly unlikely (Abrahams et al., 2006; Devine et al., 1998). Both groups, shrubs and grasses, inhabit the Jornada Experimental Range (JER). Whereas grasses dominated all soil types in the 19th century, sandy soils are currently dominated by honey mesquite (Prosopis grandulosa), while rocky soils and bajadas are dominated by creosote bush (Larrea tridentata). In both soils, grass cover is very low, usually b10% (Gibbens et al., 2005). Grass cover is higher on silty alluvial flats where they accompany tarbush (Flourensia cernua). Whereas the tarbush shrubs seem to be randomly distributed in between barren or crusted surfaces, the grasses were mainly confined to small depressions, usually 2–3 m in diameter. Moreover, grasses also occupy the under-canopy habitat of the shrubs. Given that within very small areas vegetation distribution cannot be attributed to differences in grazing pressure, studying the inter-relations of shrubs and grasses at a small research site may highlight the remaining potential factors (i.e., independent of grazing pressure) that dictate plant distribution. Since water dictates both germination (and hence cover) and biomass (Sala et al., 1988) while nutrient availability dictates mainly biomass (Gough et al., 2000) rather than germination and cover (Brown and Archer, 1989; MacMahon and Schimpf, 1981), and following lack of evidence regarding a severe shortage in nutrients in the Jornada plants (Schlesinger et al., 2006), we hypothesize that the variable distribution of grasses may stem from water availability. The current research aims therefore to evaluate the possible role of water availability in grass distribution.

2. Material and methods 2.1. The research site and previous findings A relatively flat area (bajada) sloping gently (~ 3°) westward was chosen at the Jornada Experimental Range (JER) within the northern Chihuahuan Desert, NM, US (0339313E; 3600810N) (Fig. 1). High summer temperatures and fairly low winter temperatures characterize the area, with minimal average temperatures during January of 3.8 °C and maximal average temperatures during July of 26.0 °C. Potential evaporation is high, averaging ~2200 mm (Wainwright, 2006). Average annual precipitation is 245 mm, out of which ~ 60% falls during the summer (June–September). Whereas low-intensity regional rainstorms characterize the wintertime, high-intensity, convective thunderstorms characterize the summer (Conley et al., 1992; Wainwright, 2006). The research site was characterized by Typic Calciorthid soil (approximately 25% of sand, 40% of silt and 35% of clay). The area was characterized by alternating dense macrophytes (shrubs and perennial grasses) and microphytes, i.e., biocrusts, also termed biological soil crusts. While intense grazing took place at the site until 1922, it was followed by medium-intensity grazing until 1946. Since 1947 the site experienced light grazing only (http://jornada.nmsu. edu/content/jornada-experimental-range-stocking-rates-cattle-horsesand-sheep-beginning-1916). Macrophytic patches mostly included 50–80 cm-tall and 50–80 cm-diameter tarbush (Flourensia cernua). While being often accompanied by 20–50 cm-tall tobosa grass (Pleuraphis mutica) at their under-canopy (reaching 88–95% cover), and by 10–20 cm-tall burrograss (Scleropogon brevifolius) that occupied several meter patches in between the shrubs (reaching 70–85% cover; Fig. 2a), the microphytes (biocrusts)

Fig. 1. Location of research site.

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a

b

c

d

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Fig. 2. General view of the research site showing shrubs and biocrusts surrounding grass patches at the depressions (DP) (a), lichen-dominated biocrust (CR) (b) tall tobosa grass on a small mound under tarbush (UC) (c), and bare smooth surface (PL) (d).

are comprised of high-relief lichens (Fig. 2b), with only sparse cover (b10%) of burrograss. As already previously noted (Buffington and Herbel, 1965; Devine et al., 1998; Gardner, 1951; Peters and Gibbens, 2006), all three species were often found next to one another. Unlike the shrubs that were mainly confined to isolated 5–10 cm elevated mounds (Fig. 2c), both grass types grew in dense stands (Devine et al., 1998). While confined to the under-canopy habitats in our research site, tobosagrass also occupied large sun-exposed playas, ~300 m downslope of our research site, subjected to high soil moisture content (Gallardo and Schlesinger, 1995; Huenneke and Schlesinger, 2006; Snyder et al., 2006). Bare smooth surfaces with a sparse shrub cover (10–15%) and no grass characterized the southern part of the research site (Fig. 2d). The compacted surface is underlined by a developed vesicular layer conducive for runoff generation (Blackburn, 1975). While sometimes patches of pale-green smooth and low-biomass cyanobacterial crusts can be noted, for the most parts, these surfaces lacked biocrusts and can therefore be regarded as bare surfaces. The presence of four distinct habitats in a small area suggested that water availability may dictate their biomass. Thus, due to their vesicular layer and their extremely smooth surface that facilitates runoff flow (Kidron, 2007), the bare surfaces may serve as a source of water for a large (b 100 m in diameter) playa surface 300 m downslope (as reflected by the relatively high-density gullies). We hypothesized that the lack of sufficient moisture at the bare surfaces may have lead to the absence of grass and high-biomass crusts at these surfaces (Kidron et al., 2009, 2010), and that water availability will increase from the bare surfaces to the lichen-dominated crust, the burrograss and the tobosagrass. Our hypothesis was supported by previous findings. Runoff plots that we constructed on the lichen-dominated crust and on adjacent scalped crusts at this very research site readily yielded runoff following medium and even low- to medium-intensity rains (of ≥7.5 mm h−1 lasting for ≥ 5 min). Plots in which crusts were scalped (therefore mimicking the bare surfaces) yielded approximately twice the amount of runoff than that yielded by the crusted surfaces, a phenomenon that was explained by the high microrelief of the crusted

surfaces (which acted to impede flow velocity) in comparison to the low-microrelief of the bare surfaces (Kidron et al., 2012). Yet, whether crusted or not, both surfaces exhibited high responsiveness to high-intensity rains, pointing to the potential role of the silty (loessial) parent material in contributing runoff to adjacent low-lying depressions. 2.2. Methodology In order to study the water regime at the different habitats, a 50×50 m area at the tarbush site in JER, which included all main habitats, was demarcated at the tarbush site. The habitats were: smooth silty (loessial) playa surfaces having for the most parts a bare surface (PL), surfaces having a mature lichen-dominated crust (CR), grass-covered patches (mainly burrograss) which are mainly confined in small depressions (DP), and tall grass (tobosa grass) covering the undercanopy of tarbush (UC) (Fig. 2). In contrast to the dense cover of grass at DP and UC, grass was absent at PL while being very scarce at CR. Schematic cross sections of the habitats are shown in Fig. 3. Three replicates of each of the four habitats were stratified and randomly selected for the measurements. Following similar results of moisture obtained from all ~ 12 depressions at the site (as obtained during occasional sampling in 1998), we believe that the number of repetitions suffice to adequately represent the moisture conditions of the different habitats. They included periodic moisture measurements during the summer (June–October) of 1999 and weekly measurements during the summer of 2009. Following the confinement of grasses to the upper 30 cm of the soil (Muldavin et al., 2008; Pockman and Small, 2010), soil measurements took place at a depth of 0–30 cm at 10 cm intervals. The soil samples (approx. 50 g each) were collected with an auger. The roots within the samples were sorted out rapidly in shade (within 2–3 min, thus allowing for only minimal evaporation), and the mineral soil (b2 mm) was put within flasks, which were immediately sealed. The holes were filled immediately with soil to minimize, to the best of our ability, disturbance. Nevertheless, due to technical difficulties, 7.3% of the samples had to be discarded. The samples were taken soon afterwards to a nearby lab where they were weighed, oven dried at 105 °C until reaching a constant

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Fig. 3. Schematic cross sections within the research site. While shrubs characterize elevated habitats, grass concentrates at depressions.

weight and then reweighed in order to determine their gravimetric water content. In order to determine the volumetric water content, the soil bulk density of the different depths of each habitat (N= 6) was determined. For the bulk density, soil samples that were collected in 10 cm-long and 4.75 cm-diameter tubes were oven-dried at 105 °C and weighed (Throop and Archer, 2008). Subsequently, the volumetric water content was determined following the equation: Wv ¼ Ww  Db where Wv is the volumetric water content, Ww the gravimetric water content, and Db is the soil bulk density. Following the determination of the wilting point (Kidron et al., 2002), the available water, i.e. the amount of water at field capacity subtracted by the amount of water at wilting point, was calculated, as described elsewhere (Kidron and Tal, 2012). In addition, periodic measurements of the potential evaporation took place. These were executed by 20 cm-long and 2 cm-diameter glass tubes inserted 15 cm into the ground. Since the values of PL and CR yielded insignificant differences, both habitats were combined and are referred herein as PL–CR. The tubes were protected by a 1 × 1 cm metal net from animals. The tubes were filled with deionized water up to 2 cm below their rim (to avoid water spilling following high-speed winds or high-depth rain events) and the amount lost was measured weekly using a ruler. Due to technical problems, 13.4% of the data had to be discarded. In addition, small rain gauges provided rain measurements. Grass biomass was determined within a 10.1 cm-diameter cylinders following the summer growing period. Using a 10 × 10 cm grid, 20 samples from the DP and UC were cut, oven-dried (70 °C for 48 h) and weighed. In order to measure the biomass at PL and CR, 10 transects of 10 m each were randomly established and grass that was grown at 1 m intervals within the 10.1 cm-diameter tubes was cut and weighed as described above. Two-way ANOVA was performed in order to determine the effect of habitat and time of measurement on the evaporation and the effect of the rain amount and habitat on the moisture content. In agreement with previous findings regarding the role played by >10 mm rains in runoff generation (Kidron et al., 2012), we distinguished between

>10 mm rains that fell within one week of the soil measurements and b 10 mm rains. Simple correlation analyses were performed to evaluate the relations between the average water availability and grass biomass. Differences were regarded significant at P b 0.05. 3. Results Grass biomass as measured during the end of the summer growing period of 1999 and 2009 is shown in Fig. 4. While high-density grass (70–80% cover), mainly consisting of burrograss, characterized the low depressions, tall tobosa grass with > 90% cover characterizing the relatively elevated areas at the under-canopy shrubs, as also reported elsewhere (Gardner, 1951). Above ground biomass showed significant differences (Pb 0.001) between all habitats, with UC showing the highest biomass. Annual rain amount and maximal intensities during the two summers are shown in Fig. 5. Rain amount and intensities were substantially higher during 1999 than 2009, with total summer rain reaching 235.3 and 143.0 mm during 1999 and 2009, respectively. Available water content showed pronounced differences between the habitats, with a clear gradient with PL b CR b DP b UC (Figs. 6 and 7). When a two-way ANOVA was performed, both rain amount and habitat were found to significantly affect the available water content,

Fig. 4. Average above-ground biomass at all habitats (the bare surface, PL; the biocrust, CR; the depression, DP; the under-canopy, UC) during 1999 and 2009. Bars represent 1 SE.

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Fig. 7. Average moisture content for the 0–30 cm soil depth for each habitat (the bare surface, PL; the biocrust, CR; the depression, DP; the under-canopy, UC) during 1999 and 2009. Bars represent 1 SE.

Fig. 5. Rain amount of ≥1 mm for the summers of 1999 (a) and 2009 (b) as measured at the meteorological station.

with the interaction of both variables being significant only for 1999, probably as a result of the higher runoff events during that year (Kidron et al., 2012) that enabled the differentiation between both variables (Table 1). In comparison to all other habitats, the low moisture content of PL was striking, especially due to the fact that 5 out of the 9 samplings during 1999 and 6 out of 15 samplings during 2009 were carried out within 4 days after rainstorms of ≥5 mm. However, soil moisture shortly after a high-magnitude high-intensity rain event was the highest at DP. This can be seen on the 22.7.99, 29.6.09 and 13.9.09 sampling, which was preceded respectively by 20.8, 8.0 and 12.7 mm rain events (as measured in the research site). As far as the evaporation is concerned, evaporation at UC showed substantially lower rates in comparison to DP and PL–CR, with UC showing an average evaporation rate of 4.5 mm (SE=0.2) in comparison to 5.8 mm (SE = 0.3) and 7.0 mm (SE = 0.5) at DP and PL–CR, respectively (Table 2). Indeed, when a two-way ANOVA was performed, the

habitat as well as the time of measurement was found to significantly affect the evaporation rate (Table 3). When the grass biomass was plotted against the available water content, significant relations were found. For both years, positive relations with high correlations (r 2 > 0.90) were found (Fig. 8). 4. Discussion By restricting our study to a very limited ≪ 1 ha site, we tried to avoid the impact of grazing on the current shrub–grass distribution. This is in agreement with publications that highlighted (a) the fast recovery (several years) of grass following grazing (Weaver and Albertson, 1944), (b) the similar rate of recovery following droughts for grazed and ungrazed plots (Nelson, 1934; Weaver and Albertson, 1944) and (c) findings that shrub encroachment was independent of grazing (Brown et al., 1997; Buffington and Herbel, 1965). Since grazing simultaneously affects the entire research site and following the light grazing at our research for the last ~50 years (that according to some publications may even be beneficial for grass establishment; see Paulsen and Ares, 1961 and Noy-Meir, 1974), recovery should have already taken place with plant distribution returning to the pre-grazing pattern. The current patch distribution should therefore be explained regardless of the grazing history and may therefore not reflect the grazing intensity. Confined to the summer season, during which ambient temperatures are suitable for grass activity (Herbel et al., 1972; Muldavin et al., 2008; Robertson et al., 2010), grass growth will benefit from the high summer precipitation of ~60% of the annual rain (Neave and Abrahams, 2002). With higher summer precipitation during 1999, grass biomass was indeed higher, as can be seen in Fig. 8, and in agreement with previous reports that showed a link between precipitation and inter-annual differences in grass biomass (Gibbens and Beck, 1987; Reynolds et al., 2006). Grass biomasss may also benefit from additional water such as from runoff water. Characterized by much higher intensities (Herbel et al., 1972; Reid et al., 1999), summer rains may subsequently result in high runoff

Table 1 Two-way ANOVA for available water content in relation to rain (over and below 10 mm) and habitat for 1999 and 2009. Source of variation

Fig. 6. Moisture content for the 0–30 cm deep soil layer at the four habitats (the bare surface, PL; the biocrust, CR; the depression, DP; the under-canopy, UC) as obtained during the summer of 1999 (a) and 2009 (b). Arrows and values indicate the rain events and amounts as measured at the site. Bars represent 1 SE.

Rain Habitat Rain × Habitat Error Total

SS

df

F

P-value

1999

2009

1999

2009

1999

2009

1999

2009

1155.3 1530.7 377.2 705.4 8331.9

295.1 1314.8 27.4 865.3 6518.3

1 3 3 97 105

1 3 3 126 134

158.9 70.2 17.3

43.0 63.8 1.3

b0.001 b0.001 b0.001

b0.001 b0.001 0.268

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Table 2 Periodic measurements of potential evaporation at PL–CR, DP and UC during the summer of 2009. One standard errors are in parentheses. Date

PL–CR

DP

UC

23.6

49.8 (0.2) 57.8 (0.2) 69.2 (2.8) 62.9 (1.9) 48.5 (6.1) 44.8 (2.2) 48.4 (2.4) 43.5 (2.1) 54.8 (0.5) 37.3 (2.2) 49.3 (5.2) 36.2 (0.7) 34.8 (0.4) 48.9 (2.8) 7.0 (0.5)

32.9 (1.4) 39.8 (1.4) 63.7 (3.1) 53.8 (2.4) 39.0 (4.4) 42.3 (2.6) 45.0 (1.6) 36.5 (0.3) 49.3 (1.8) 28.2 (1.2) 45.3 (1.9) 31.4 (3.9) 31.8 (1.5) 41.6 (2.8) 5.8 (0.3)

23.5 (2.5) 29.0

28.6 8.7 19.7 26.7 2.8 9.8 16.8 26.8 30.8 6.9 13.9 20.9 Average Average per day

45.2 (6.5) 47.5 (4.2) 27.3 (4.1) 33.7 (3.5) 29.3 31.0 (2.0) 43.8 (1.9) 22.8 (1.8) 35.3 (2.6) 25.8 (0.7) 26.2 (2.7) 32.9 (2.3) 4.5 (0.2)

generation (Kidron et al., 2012; Parsons et al., 2003; Reynolds et al., 2006). A continuous and relatively dense network of gullies at CR and especially at PL carries the runoff water to large (> 100 m diameter) depressions (playas), some of which remain flooded for > 1 month. At our site, the relatively smooth surface of CR provides runoff water to the small depressions at DP, which results in turn in higher moisture content and extended wetness duration. In fact, moisture at DP was the highest shortly after runoff events as can be seen during 22.7.99, 29.6.09 and 13.9.09 (Fig. 6). Attesting to run-on zones, these wet surfaces were consistent with the areas inhabited by dense grass stands at DP (Fig. 9). While high grass cover may act to deplete the soil moisture by transpiration (Gutschick and Snyder, 2006), it may, on the other hand, also act to shade the surface from direct irradiance, and thus act to impede potential evaporation (Table 2). In agreement with other reports at the southwestern USA (Kieft et al., 1998; Patten, 1978; Scholes and Archer, 1997) and other arid and semi-arid regions (Cantón et al., 2004; Cerdà, 1997; Pariente, 2002; Schade et al., 2003), high moisture content characterized UC, as also found for JER (Grover and Musick, 1990; Hennessy et al., 1985; Snyder et al., 2006). While runoff addition was often seen as the causal factor for the high moisture content at UC (Cantón et al., 2004; Puigdefábregas, 2005), runoff could not have been the contributing factor for the high UC moisture content at our site. With most under-canopies having low mounds, runoff did not concentrate at the under-canopies of the shrubs, in agreement with similar findings

Fig. 8. The relationships between soil moisture content and grass biomass in the four habitats. Bars represent 1 SE.

from the Nizzana research site, in the western Negev Desert, Israel (Kidron, 2010, 2011). The UC also did not contribute runoff water as verified during high-intensity rains in the field. Similar to previous observations (Abrahams et al., 2006; Devine et al., 1998; Reid et al., 1999), this may be explained by the efficiency in which the dense stand of grass impedes runoff flow. While not explained by differences in the interrelations between input and output, the high moisture content at UC could have been explained by the low evaporation rates that characterized UC. As UC is shaded from direct sun beams, evaporation rates, as periodically measured at UC, were approximately 0.67 that of the sun-exposed habitats CR and PL (Table 2). The differences were similar to three year-long measurements at the Nizzana research site at the Negev Desert during which evaporation at UC was 0.53 that of the adjacent sun-exposed habitat (Kidron, 2009). In this regard it is worth mentioning that an improvement of the physical conditions due to high soil organic matter (SOM) (Armas and Pugnaire, 2005; Pugnaire et al., 2004) and the subsequent increase of the water-holding capacity (Boix-Fayos et al., 2001; Joffre and Rambal, 1988) may also contribute to the higher moisture content at UC. The relatively high moisture content recorded at UC was in agreement with other data reported from the southwestern USA (Hennessy et al., 1985; Kieft et al., 1998; Patten, 1978; Scholes and Archer, 1997; Snyder et al., 2006). There was no indication in any of these cases of the involvement of hydraulic lift (Caldwell et al., 1998; Richards and Caldwell, 1987). Furthermore, if water was to be hydraulically redistributed during summertime it would take place from the upper wet surface to depth (Burgess et al., 1998; Ryel et al., 2003), and thus would deplete rather than contribute to the moisture content at the 0–30 cm soil layer of UC.

Table 3 Two-way ANOVA for evaporation in relation to habitat and time of measurement. Source of variation

SS

df

F

P-value

Time of measurement Habitat Time × Habitat Error Total

7937.8 4546.7 785.9 1399.4 197,291.0

12 2 24 67 106

31.7 108.8 1.6

b0.0001 b0.0001 0.077

Fig. 9. Longer duration of moisture characterizes the topographically low areas (run-on zones) with burrograss, as can be seen during 3.8.99, approximately 24 h after a runoff-induced 7 mm rain event. Note the link between these areas to grass distribution.

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While UC and DP exhibited relatively high water content, PL and CR exhibited relatively low water content, which also implies shorter growth duration. In this regard, a comparison to the Negev Desert may be of interest. Water at UC was available for ~1 month longer than at the adjacent exposed habitat, i.e., extending the growing period for up to ~ 25% (Kidron, 2010). Similarly, higher cover characterized the annuals at the shrub's under-canopies at the Negev (Holzapfel et al., 2006), as was also the case in other arid and semi-arid areas (Armas and Pugnaire, 2005; Holzapfel and Mahall, 1999; Maestre et al., 2003). The higher cover of annuals at the under-canopies in the Negev could not have been attributed to the SOM content (Kidron, 2010). This is in agreement with research in the US that failed yet to show a close link between nutrient amounts and annual or grass germination (Beatley, 1969; Brown and Archer, 1989; Grover and Musick, 1990; Went, 1949, 1973). With water being the main limiting factor in arid zones (Noy-Meir, 1973) and principally dictating germination (Brown and Archer, 1989; Grover and Musick, 1990), competition for water is expected. Yet, we argue that temporal and spatial factors may minimize the competition between shrubs and grasses. Since runoff is almost solely limited to summer rains while shrub seedlings germinate after winter rains (Brown and Archer, 1990; Brown et al., 1997), shrub seedlings would not be confined to run-on zones. Shrubs may utilize the large, long-lasting (Muldavin et al., 2008), and subsequently deep-infiltration winter rains (Wilcox et al., 1997), thus reaching beyond the 20–30 cm depth utilized by the grass (Brown and Archer, 1990; Herbel et al., 1972; Kincaid et al., 1964; Walter, 1979). As this is taking place at the time during which grasses are still dormant or not yet fully active (Herbel et al., 1972; Snyder et al., 2006), one may thus expect competition between the shrub seedling and the grasses to be minimal (Pockman and Small, 2010; Walter, 1979). Temporal variability (time of activity) thus leads to spatial variability, either laterally (by controlling the distributional patterns in accordance with run-on zones), or vertically (by the partition of water utilization from shallow or deep soil layers). Root segregation (Alados et al., 2006; Armas and Pugnaire, 2005, 2009; Soliveres et al., 2010) may thus lead to niche partitioning between shrubs and grasses (Briones et al., 1998; Gebauer et al., 2002; Sala et al., 1989) during which C3 tarbush is providing the C4 tobosa grass with a preferentially shaded habitat. Root segregation facilitates the growth of a dense stand of tobosa grass under the shrubs, similarly to the dense stands of tobosa grass at the relatively wet habitats of the large playas. Since mesic habitats are likely to support higher biomass (whether of high-biomass individuals of a given species or the colonization of tall-morphology plants), we interpret the UC support of tobosa grass (which is characterized by higher density and biomass) as facilitation. This is in agreement with other publications that proposed that an association in which one organism (the tarbush in our case) increases the distribution of another organism (tobosa grass) beyond the habitats which they normally occupy (large playas) may be regarded as facilitation (Bruno et al., 2003). And thus, as also found by Báez and Collins (2008) for the Chihuahuan Desert and Holzapfel et al. (2006) for the shrub–annual interrelations in the Negev, and in agreement with other findings concerning shrub–grass interactions (Knoop and Walker, 1985; Scholes and Archer, 1997; Walker and Noy-Meir, 1982), we therefore conclude that the shrub–grass association in our research site may be better described as facilitation rather than competition. Yet, further research is called for in order to assess whether the interaction is species-specific (Callaway, 1998). Since soil moisture rather than precipitation per se mainly determines ecosystem productivity (Huenneke and Noble, 1996; Muldavin et al., 2008; Noy-Meir, 1973; Pockman and Small, 2010; Reynolds et al., 2006; Robertson et al., 2009, 2010), and following the (a) positive role played by water redistribution in deserts (Noy-Meir, 1973; Reid et al., 1999), and (b) the quick and positive response of grass to water addition (Robertson et al., 2010; Throop et al., 2012), the higher water content measured at UC and DP may point to the cardinal role played

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by water in the formation of “islands of fertility” and in shrub–grass association. Thus, in line with other reports that highlight the importance of runoff (Noy-Meir, 1986; Rango et al., 2006) and of impeded evaporation (Noy-Meir, 1973; Herbel and Gile, 1973; Campbell and Harris, 1977: Scholes and Archer, 1997) to plant and grass establishment, our findings point to the importance of water availability for grass persistence and its role as a key driver for plant productivity (Muldavin et al., 2008; Pockman and Small, 2010). Our findings suggest that the decrease in grass cover in JER may stem from a shortage of sufficient soil moisture during summertime, an expected outcome of global warming. Located at the desert fringe and at near threshold conditions for grass (Conley et al., 1992; Gardner, 1951; Shreve, 1917), the northern Chihuahuan Desert may thus be especially vulnerable (Monger, 2003), and among the first areas to experience the first impact of global warming. Acknowledgments The project was supported by grant #00R-009 of the International Arid Land Consortium (IALC). We would like to thank H. Curtis Monger (New Mexico State University) for his valuable support during the initial phase of the research and Abraham Starinsky (Hebrew University) for his valuable help in acquiring the funds, Tim Jones from NMSU, John Anderson from the Jornada Long-Term Ecological Research Program, Robin Foldager, Adrian Godina, Randy De La O, Amy Slaughter and the staff at the USDA-Jornada Experimental Range for their assistance with the study, Jeanne Tenorio for her devotion and valuable assistance in the field and Carol A. Kidron for the editing. The valuable comments made by Curtis Monger on an earlier version of the manuscript and the valuable comments made by two anonymous reviewers are highly appreciated. References Abrahams, A.D., Neave, M., Schlesinger, W.H., Wainwright, J., Howes, D.A., Parsons, A.J., 2006. Biochemical fluxes across piedmont slopes of the Jornada basin. In: Havstad, K.M., Huenneke, L.F., Schlesinger, W.H. (Eds.), Structure and Function of a Chihuahuan Desert Ecosystem. The Jornada Basin Long-Term Ecological Research Site. Oxford University Press, NY, pp. 150–175. Alados, C.L., Gotor, P., Ballester, P., Navas, D., Escos, J.M., Navarro, T., Cabezudo, B., 2006. Association between competition and facilitation processes and vegetation spatial patterns in alpha steppes. Biological Journal of the Linnean Society 87, 103–113. Armas, C., Pugnaire, F.I., 2005. Plant interactions govern population dynamics in a semi-arid plant community. Journal of Ecology 93, 978–989. Armas, C., Pugnaire, F.I., 2009. Ontogenetic shifts in interactions of two dominant shrub species in a semi-arid coastal sand dune system. Journal of Vegetation Science 20, 535–546. Báez, S., Collins, S.L., 2008. Shrub invasion decreases diversity and alters community stability in northern Chihuahuan Desert plant communities. PloS One 3, e2332. Beatley, J.C., 1969. Biomass of desert winter annual plant populations in southern Nevada. Oikos 20, 261–273. Bestelmeyer, B.T., Brown, J.R., Havstad, K.M., Fredrickson, E.L., 2006. A holistic view of an arid ecosystem: a synthesis of research and its applications. In: Havstad, K.M., Huenneke, L.F., Schlesinger, W.H. (Eds.), Structure and Function of a Chihuahuan Desert Ecosystem. The Jornada Basin Long-Term Ecological Research Site. Oxford University Press, NY, pp. 354–468. Blackburn, W.H., 1975. Factors influencing infiltration and sediment production of semi-arid rangelands in Nevada. Water Resources Research 11, 929–937. Boix-Fayos, C., Calvo-Cases, A., Imeson, A.C., Soriano-Soto, M.D., 2001. Influence of soil properties on the aggregation of some Mediterranean soils and the use of aggregate size and stability as land degradation indicators. Catena 44, 47–67. Briones, O., Montaña, C., Ezcurra, E., 1998. Competition intensity as a function of resource availability in semiarid ecosystems. Oecologia 116, 365–372. 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., 1990. Water relations of a perennial grass and seedling vs adult woody plants in a subtropical savanna, Texas. Oikos 57, 366–374. Brown, J.H., Valone, T.J., Curtin, C.G., 1997. Reorganization of an arid ecosystem in response to recent climate change. PNAS (USA) 94, 9729–9733. Bruno, J.F., Stachowicz, J.J., Bertness, M.D., 2003. Inclusion of facilitation into ecological theory. Tree 18, 119–125. Buffington, L.C., Herbel, C.H., 1965. Vegetational changes on semidesert grassland range from 1958 to 1963. Ecological Monographs 35, 139–164. Burgess, S.S.O., Adams, M.A., Turner, N.C., Ong, C.K., 1998. The redistribution of soil water by tree root systems. Oecologia 115, 306–311.

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