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Ecological dynamics in plant communities determine their natural changes in composition, structure, and shape (Vilanova et al., 2006). Overgrazing in arid rangelands is an important disturbance in such fragile ecosystems (Belsky, 1992; Valone, 2003; Higgins et al., 2007; Distel, 2013). Other disturbance factors that can alter natural ecosystem dynamics are human activities, such as land use changes (Zhao et al., 2015) and introduction of invasive species (Burgiel and Muir, 2010). Social and economic incentives have indeed promoted agriculture and natural habitat-related changes through natural resource management (Gutierrez-Cedillo et al., 2008). Similarly, changes in structure and density of plant communities through modifications of the intensity and frequency of grazing for the purpose of increasing species productivity have also been frequently induced (Gillen et al., 1991; Craig, 1999; Fuhlendorf et al., 2010; Catorci et al., 2011; Vermeire et al., 2014).

Herbivores are important drivers of change in plant communities (Olff and Ritchie, 1998; Su et al., 2015). Kimball (1980) found that the pressure caused by consumption of some shrub species affects plant size and the formation of new leaves and branches, consequently affecting photosynthesis (Seagle and Liang, 2001; Nabity et al., 2009). At a different scale, composition may be affected when species that are palatable to herbivores become scarce through browsing and trampling (Milchunas and Lauenroth, 1993; Fleischner, 1994; Lauenroth et al., 1999). The effects of grazing by bovines have been widely studied (Noss, 1994; Bowman et al., 2008). Trampling and grazing by livestock increases soil loss and modifies vegetation structure as well (DeLuca et al., 1998; Menard et al., 2002). In both cases, large herbivores modify the habitat they inhabit. Plant species selection and its impact on vegetation may change depending on grazing space available. The previous process applies for cattle, horses, and donkeys; the herbaceous strata is one of the most affected due to large herbivore feeding preferences, consisting of grasses (Reiner and Urness, 1982; Menard et al., 2002).

In Mexico, studies of the impacts caused by free grazing on rangelands have mainly focused on topics such as land tenure, communal organization and overgrazing (Manzano et al., 2000), loss of native species, and desertification (Ceballos et al., 2010). A growing interest has emerged in using rangeland recovery as a conservation tool to increase productivity, provide environmental services and habitat for wildlife (Ceballos et al., 2010), and buffer the adverse effects of drought and climate change (Vogel et al., 2012). In Coahuila, because of overgrazing pressures, a reduction in livestock was promoted. Consequently, between 1991 and 2007 livestock numbers decreased by approximately 11.2% for cattle and 42.6% for goats, thus reducing surface covered by rangelands by 10.6%. The latest information from the national livestock census (Instituto Nacional de Estadistica, Geografia e Informatica, coah) showed that Coahuila was still the twelfth Mexican state in equine abundance with approximately 53,700 horses, and 15,000 donkeys. Recent estimates of cattle and equines including cows, heifers, and stallions for Coahuila are as large as 735,000 (Secretaria de Agricultura, Ganaderia, Pesca y AlimentacicSn, Considering the potential effects of these animals on vegetation and that desert microphyllous and rosetophyllous scrublands are the most widely distributed natural plant communities in the Chihuahuan desert, we aimed at determining livestock impact on plant species diversity and vegetation structure in these vegetation communities. Even though structural changes are often visible, there are insufficient ecological references about the changes in these types of scrublands in the Mexican part of the Big Bend-Rio Bravo region. The previous region represents a huge transboundary system of conservation area. Therefore, current references regarding this topic are of great importance for conservation and management. We hypothesized that livestock grazing would affect floristic composition and vegetation structure in both microphyllous and rosetophyllous scrublands.

Materials and Methods--Our study was conducted in northwestern Coahuila, Mexico (Fig. 1) at the municipality of Ocampo (29[degrees]01'N, 102[degrees]54'W) where xeric scrublands are the dominant vegetation types, followed by grasslands (Villarreal-Quintanilla and Valdes-Reyna, 1993; Powell, 1997). The main regional climate type is dry, and the climate subtype is very dry-hot with very occasional rains throughout the year (the climatic formula is BWh; Peel et al., 2007). According to data from the National Meteorological Service, mean annual temperature in this region is 21.2[degrees]C, with mean low of 12.5[degrees]C and a high of 29[degrees]C. Mean annual rainfall is 236 mm, and on average, the last 60 years had 28.5 days of rain each year. Elevation ranges from 560 to 940 m.

Field Evaluation--The area subject to livestock presence comprises a 68,100-ha polygon in lands within the ejidos of Jaboncillos Grande, San Vicente y Zacatonal, La Union y El Olan, and Ojo Caliente. These rural communities are located at the municipality of Ocampo, Coahuila, in northern Mexico. The area without livestock was located at El Carmen Ranch (13,740 ha). This ranch has been managed for conservation. Consequently, livestock has been absent for the last 15 years. We assessed the effect of grazing on vegetation by comparing mean values of plant density, frequency, and cover in sites with and without grazing. We detected the presence of grazing by direct observation of animals, evidence of browsing, and presence of feces. We surveyed the rural communities of the study area, and estimated around 1,500 animals in free grazing (60% cattle, 40% equines; J. J. Ochoa, pers. observ.). We considered sites without livestock grazing as our controls. We also assessed seasonal effects by making comparisons between the dry (January-July) and wet (August-December) seasons.

We established sampling plots in two different plant communities: desert microphyllous and desert rosetophyllous scrublands. We randomly selected sampling plots within these vegetation communities based on presence of livestock. The minimum distance between sampling plots was 1 km, with the maximum distance not exceeding 5 km. Thirty-six sampling plots (76.5%) were managed with livestock, whereas 11 sampling plots (23.5%) had no livestock. However, other wild, nondomestic ungulates such as mule deer (Odocoileus hemionus), white-tailed deer (Odocoileus virginianus), and collared peccary (Tajasu tajacu)--which are the largest wild herbivores naturally occurring in the region--(McKinney and Delgadillo Villalobos, 2014) have been present in these control sites since livestock exclosures were initiated 15 years ago.

We sampled vegetation from June 2011 to August 2013. Sampling took place both during the dry (January-July) and wet (August-December) seasons. We sampled a total of 23 sampling plots in the dry season, and 24 sites in the wet season in order to relate season with vegetation parameters. From all 47 sampling sites, 19 correspond to microphyllous scrublands and 28 to rosetophyllous scrublands. We calculated the number of sampling plots based on cover proportion and its standard deviation obtained from each vegetation type based on a prospective sampling. Our sampling design included replicates for all management-plant community-season combinations.

To determine plant cover, we sampled five 10-m Canfield lines (Coulloudon et al., 1999) in each sampling plot. We established two 10-m Canfield lines facing north, starting from one center reference point, and then the other three lines were located at each cardinal direction, in the shape of a cross. To quantify plant species richness, abundance, dominance, and frequency in three vegetation height strata, we established one rectangular (10 x 5-m) plot in each sampling plot; such plots were always located parallel to the north Canfield line. Within each plot we counted and recorded 1) all species higher than 1.5 m, 2) all shrub species (between 0.5 and 1.5 m in height), and 3) herbaceous species (<0.5m in height) within a 5 x 5-m net quadrant established inside the 10 x 5-m rectangular sample plots. We recorded these herbaceous species inside 1 x 1-m plots located in opposite corners of the 5 x 5-m sample plots (Elzinga et al., 1998).

Using the procedure described by Canizales-Velazquez et al. (2009), we estimated ecological parameters including relative abundance, relative dominance, relative frequency of plant communities, and the importance value index (IVI), which is the sum of these three parameters. Similarly, we estimated species richness, the potential number of species, sampling efficiency, plant diversity, and similarity as a measure of ecological composition. For all five estimates, abundance was used as a measure of reference for the calculations.

Data Analysis--We used a Student t test for independent samples (Daniel, 2005) to evaluate differences in means of the IVI, density, frequency, and coverage between sites with and without livestock grazing. To estimate the ecological variables, we used abundance, i.e., the number of individuals per plant species, recorded at each condition (with and without livestock grazing), season, and for each vegetation type. We also calculated plant richness, in each of the two conditions analyzed, and in each season. To estimate potential number of species we used the Chao 1 and Jackknife 1 nonparametric estimators. We chose these estimators because 1) we did not assume a previous abundance distribution model, 2) they are robust when calculating minimum estimates of species richness, 3) their use is recommended as a recurrent measure in analyses of biodiversity, 4) Chao 1 is based on abundance data, or singletons and doubletons, and Jackknife 1 (incidence) is based on unique species, or species found in only one sample (Magurran, 2004; Hortal et al., 2006), and 5) Jackknife indices tend to be a conservative estimator. The use of both Chao 1 and Jackknife 1 may provide an estimated range of species richness (Villarreal et al., 2006). We calculated the estimators through 100 randomizations using the software EstimateS, version 8.2 (statistical estimation of species richness and shared species from samples;, based on the number and abundance of species found per sampling unit (Canfield lines). We also measured sampling efficiency through the Clench model, the coefficient of determination ([R.sup.2]), and the slope of the species accumulation curve, which evaluates inventory quality. Their calculation (in EstimateS) was based on the number of samples (i.e., Canfield lines) in each type of scrubland, condition of presence or absence of livestock grazing, and season.

After testing for normality of the data, we conducted a one-way analysis of variance and a post hoc Tukey test with logarithmic transformed data to compare the effect of livestock on both types of desert scrublands. We used software Statistica version 8.0 (data analysis software system; to perform this statistical analysis. We also evaluated alpha and beta diversity for each condition and season; the Simpson diversity index (1--1/D) and the Shannon diversity index (H') (Magurran, 2004) were used as an alpha diversity measure, whereas the Bray-Curtis similarity index (Sorensen's quantitative index; Magurran, 2004) was used as a measure of beta diversity, to evaluate effects of vegetation type, condition, and season.

Results--Composition--For the microphyllous desert scrubland, we obtained a total of 1,378 plant records belonging to 24 families, 53 genera, and 66 species (Table 1). Agavaceae was the most abundant family with 431 records, representing 31.3% of total abundance in this scrubland type. The Poaceae family had the highest species richness with 19.7% of all species. We recorded 38 plant species exclusively in areas with livestock, including some opportunistic nonpalatable weeds such as twinleaf senna (Senna bauhinioides) and Fendler's bladderpod (Lesquerella fendleri), and some invasive shrubs including tree tobacco (Nicotiana glauca) and white thorn acacia (Acacia constricta).The inventory quality assessment through the Chao 1 and Jacknife 1 estimators showed that the 66 species registered in microphyllous scrub represent 80.8-88.4% of the total richness for this plant community. The data also showed reasonable adjustment to the Clench model ([R.sup.2] = 0.99). Regarding the comparison between livestock presence or absence, microphyllous scrublands registered 48% similarity, and between seasons the Bray-Curtis index yielded 75% similarity.

The rosetophyllous scrub had 2,919 plant records belonging to 25 families, 60 genera, and 75 species. As in the microphyllous scrub, Agavaceae was the most abundant family with 33.9% (990 records) of the total abundance. Moreover, Poaceae was the family with the highest species richness, with 17.3% of the total recorded species. For the livestock presence vs. absence treatment comparison, we recorded significant differences (P [less than or equal to] 0.05) in number of plant species (Table 1). Whereas for the presence-of-livestock treatment the recorded number species represents 77.0-83.5% of the estimated richness, for the livestock absence scenario the 56 identified species comprise 76.3-78.7% of the estimated richness. The Clench model yielded 86-92 potential species. Therefore, the total number of identified species represent 81.8-86.8 of the total richness. The Bray-Curtis similarity index showed no significant differences in composition related to livestock presence in the rosetophyllous scrublands (65% similarity). Species composition showed no differences between seasons (80% similarity).

We registered higher abundance and higher richness in the rosetophyllous scrubland than in the microphyllous scrubland; however, the Shannon's biodiversity index was lower in rosetophyllous (2.81) than in microphyllous scrubland (2.98). Similarity between both scrubland types was 62%.

Structure--For microphyllous scrub and areas with livestock, we registered a high dominance of creosote bush (Larrea tridentata; 16.6%), and blue grama (Bouteloua gracilis; 15.6%). We found seven plant species that occurred exclusively in areas without livestock; these included scarlet hedgehog cactus (Echinocereus coccineus), West Indian shrub verbena (Lantana urticoides), bicolor fanmustard (Nerisyrenia camporum), purple threeawn (Aristida purpurea), lovegrass (Eragrostis species), tanglehead (Heteropogon contortus), and alkali sacaton (Sporobolus airoides). The species with the highest IVIs for microphyllous scrubland were L. tridentata, Agave lechuguilla, and Opuntia leptocaulis; however, only Prosopis glandulosa showed significant differences in IVI between sites with and without livestock (Table 2). Agave lechuguilla, L. tridentata, and Parthenium incanum showed the highest IVIs in the rosetophyllous scrub, only L. tridentata was significantly different in IVI between sites with and without livestock (Table 2).

The differences between microphyllous and rosteophyllous scrublands were remarkably higher in the wet season. The sites with livestock had higher IVIs than sites without grazing, and the average IVI was higher in microphyllous than in rosetophyllous scrubland (31.43 vs. 23.17, respectively).

We observed that the zones with microphyllous scrubland without livestock grazing had the highest IVIs in herbaceous strata compared with the places under livestock use (Fig. 2). The common species were black grama (Bouteloua eriopoda) and blue grama, the latter having the highest IVI for both types of shrubland when livestock was absent (34.2 and 28.3 for microphyllous and rosetophyllous, respectively), whereas the average IVI for both types of management was not different in rosetophyllous scrubland (13.6 without vs. 14.4 with livestock), and seven species of grasses were persistent, no matter the type of management. In the scrubland strata, microphyllous had the highest IVI in livestock sites, with the same tendency in rosetophyllous, but with no statistical difference in the IVI (Fig. 2).

Discussion--In this study, we found differences in plant diversity and vegetation structure of microphyllous and rosetophyllous desert scrublands. We registered 50 shrub species, representing more than 10% of all shrub species described for this region in the floristic inventory made by Villarreal-Quintanilla and Valdes-Reyna (1993). In Brazil, Bridgewater et al. (2004) recorded 120 plant species, which represented less than 15% of all plant species in the community and those species contributed as much as 75% to the IVI. Estrada-CastillcSn et al. (2011) found that 39 of all 233 plant species recorded in Piedmont scrubland vegetation of northeastern Mexico comprised 95% of the physiognomy of this plant association. Given such information, we can conclude that a relative low number of species may very well determine the prevailing trend for this plant community.

Rosetophyllous scrubland with livestock had the highest number (68) of plant species. Disturbances influence species richness at different levels; for example, in a study developed in an arid environment in South Africa, Todd (2015) found that plant richness increased when sources of water for livestock were more distant. Beever et al. (2008) found up to 12 more species in areas under horse grazing in the Great Basin. This is consistent with other authors who found that heavier grazing intensity was associated with higher species richness (De Bello et al., 2007; Renne and Tracy, 2007; Alanis Rodriguez et al., 2008). In contrast, Al-Rowaily et al. (2015) found higher richness in sites without grazing. In our study the sites without livestock had fewer plant species, and the time of exclusion of those sites could influence richness given the competitive exclusion principle (Valone, 2003) because grazing disturbance is almost absent. The combination of two types of disturbances might also influence richness; such was the case for fire and grazing (Belsky, 1992; Valone, 2003).

Hernandez et al. (2000) studied three types of forests in central Mexico assessing cattle effects on plant species richness and vegetation structure, and found differences in plant species composition between vegetation types, but not between sites with and without cattle. They also found that two species of shrubs, capulincillo (Parathesis villosa) and palo canelo (Cornus disciflora), had the highest IVIs. We found two species of shrubs (L. tridentata and A. lechuguilla) with the highest IVIs for both types of scrubland. We found two weedy species, S. bauhinioides and L. fendleri, in the livestock sites; they are considered to be overgrazing indicators. This finding is also consistent with those from Gillen et al. (1991) and Renne and Tracy (2007) who found that weeds increased in sites under grazing pressure. Ralphs et al. (1990) detected that grazing caused substitution of desirable grasses by toxic plants, affecting abundance, dominance, plant frequency, and thus overall plant species composition. An opposing idea presented by Beever et al. (2008) indicated that sites with wild horses had more grazing-resistant shrubs and exotic plant species. Other studies (Yancey and Douglas, 1983) reported results similar to ours in terms of the influence of big herbivores like horses in plant composition in desert ecosystems. In our study, B. gracilis, which showed more than 15% dominance in rosetophyllous scrubland, stands out as having good forage value (Gillen et al., 1991).

We observed for both vegetation communities that sites with domestic herbivores had more plant cover, frequency, and abundance, regardless of season. Perhaps historical use of rangelands could help explain increases or decreases in plant species (Milchunas and Lauenroth, 1993; Fleischner, 1994). Scrub species such as honey mesquite, creosote bush, and American tarwort (Flourensia cernua) had the highest importance values in sites occupied by cattle and horses, thus influencing the structure and species composition of the plant community (DeLuca et al., 1998; Menard et al., 2002). However, comparing the vegetation structural values between shrub and herbaceous strata, we noticed that herbaceous species were higher in cover and frequency when livestock were absent, suggesting a more intense use of grasses in sites with livestock. Such differences were not so evident because the shrubs were the dominant strata. In Utah deserts, Reiner and Urness (1982) studied the diet selectivity of wild horses and found that selectivity influences plant species composition, IVI, and vegetation structure. They found that horses avoid feeding on antelope bitterbrush (Purshia tridentata), a plant species highly used by wildlife, and 40-86% of their diet was composed of grasses, followed by forbs and shrubs in no more than 5%. In our study, although diet was not analyzed, grass species like blue grama and black grama had a high IVI in sites not occupied by livestock. This result illustrates the level of use compared with sites with grazing pressure as well as the percentage of cover they represent in such sites. Beever et al. (2008), however, using a methodology that is very similar to ours, found that sites without wild horses had more shrub coverage, and two to three times more grass cover. Davies et al. (2014) did not find differences in herbaceous cover and density between sites with and without wild horses with the exception of sites with very high grazing pressure, where the cover of native grasses was reduced. There were also desirable species with high forage value that were completely replaced, such as chino grama, which was absent in the grazed sites.

There are different opinions about the influence of livestock grazing on vegetation. One of these is that the impact cannot be detected by measuring the presence of palatable plants or that at the landscape level there are other variables such as drought that influence normalized difference vegetation values. We worked on a field-level scale, with different time-space comparison scenarios, which made change detection difficult. Nevertheless we analyzed vegetation structure values as indicators of change and found that cover values did not differ between the livestock presence-absence scenarios. Instead, species abundance was higher in both types of scrublands when livestock grazing was present.

In general, the structure of both types of scrubland showed changes, mostly in species abundance, when livestock occupied the rangeland. Plant species composition in rosetophyllous and microphyllous scrublands in northwestern Coahuila had changes as well when livestock was present, especially in herbaceous strata, given that some good forage-value species like chino grama were not found in livestock sites, while other plants considered weeds or overgrazing indicators were only found in places with grazing (e.g., S. bahuinioides). Changes in composition were more visible between scrubland types compared with herbaceous and tree strata. This is evident because IVI was higher for a common group of plants between two types of scrubland.

Ecological knowledge about xeric scrublands in an arid environment, which are used for livestock raising, is useful for management decisions on 1) range management, by means of the recognition of desirable species under high grazing pressure, i.e., which seeds could be used for grassland rehabilitation; 2) conservation, using IVI as a referential value in restoration projects; and 3) future scientific research, given the lack of floristic inventories, or other related information in the region, where we documented under a scientific framework the impacts caused by activities such as livestock raising on rangelands and under xeric plant communities.

Literature Cited

Alanis Rodriguez, E., J. Jimenez Perez, O. Aguirre Calderon, E. Trevino Garza, E. Jurado Ybarra, and M. Gonzalez Tagle. 2008. Efecto del uso del suelo en la fitodiversidad del matorral espinoso tamaulipeco. Ciencia UANL XI 1:56-62.

Al-Rowaily, S. L., M. I. El-Bana, D. A. Al-Bakre, A. M. Assaeed, A. K. Hegazy, and M. B. Ali. 2015. Effects of open grazing and livestock exclusion on floristic composition and diversity in natural ecosystem of western Saudi Arabia. Journal of Biological Sciences 22:430-437.

Beever, E. A., R. J. Tausch, and W. E. Thogmartin. 2008. Multiscale responses of vegetation to removal of horse grazing from Great Basin, USA mountain ranges. Plant Ecology 196:163-184.

Belsky, J. 1992. Effects of grazing, competition, disturbance and fire on species composition and diversity of grassland communities. Journal of Vegetation Science 3:187-200.

Bowman, D. M.J. S., J. E. Riley, G. S. Boggs, C. E. R. Lehmann, and L. D. Prior. 2008. Do feral buffalo (Bubalus bubalis) explain the increase of woody cover in savannas of Kakadu National Park, Australia? Journal of Biogeography 35:1976-1988.

Bridgewater, S., J. A. Ratter, and J. F. Ribeiro. 2004. Biogeographic patterns, B-diversity and dominance in the Cerrado biome of Brazil. Biodiversity and Conservation 13:2295-2318.

Burgiel, S. W., and A. A. Muir. 2010. Invasive species, climate change and ecosystem-based adaptation: addressing multiple drivers of global change. Global Invasive Species Programme (GISP), Washington, D.C., and Nairobi, Kenya.

Canizales-Velazquez, P. A., E. Alanis-Rodriguez, R. Aranda-Ramos, J. M. Mata-Balderas, J. Jimenez-Perez., G. Alanis-Flores, J. I. Uvalle-Sauceda, and M. G. Ruiz-Bautista. 2009. Caracterizacion estructural del matorral submontano de la Sierra Madre Oriental, Nuevo Leon, Mexico. Revista Chapingo Serie Ciencias Forestales y del Ambiente 15:115-120.

Catorci, A., G. Ottaviani, I. Vitasovic Kosic, and S. Cesaretti. 2011. Effect of spatial and temporal patterns of stress and disturbance intensities in a sub-Mediterranean grassland. Plant Biosystems-An International Journal Dealing with all Aspects of Plant Biology 1-16. doi:10.1080/11263504.2011.623192.

Ceballos, G., A. Davidson, R. List, J. Pacheco, P. Manzano-Fischer, G. Santos-Barrera, and J.Cruzado. 2010. Rapid decline of a grassland system and its ecological and conservation implications. PLoS ONE 5: e8562. doi:10.1371/journal.pone.0008562.

Coulloudon, B., K. Eshelman, J. Gianola, N. Habich, L. Hughes, C. Johnson, M. Pellant, P. Podborny. A. Rasmussen, B. Robles, P. Shaver, J. Spehar, and J. Willoughby. 1999. Sampling vegetation attributes. Interagency Technical Reference 17344. Report BLM/RS/ST-96/002+1730. Bureau of Land Management, Department of Interior, Denver, Colorado.

Craig, A. B. 1999. Fire management of rangelands in the Kimbereley low-rainfall zone: a review. Rangeland Journal 21:39-70.

Daniel, W. W. 2005. Bioestadistica: base para el analisis de las ciencias de la salud. Fourth edition. Limusa Wiley, Mexico, D.F.

Davies, K. W., G. Collins, and C. S. Boyd. 2014. Effects of feral free-roaming horses on semi-arid rangeland ecosystems: an example from the sagebrush steppe. Ecosphere 5:1-14.

De Bello, F., J. Leps, and M. T. Sebastia. 2007. Grazing effects on the species-area relationship: variation along a climatic gradient in NE Spain. Journal of Vegetation Science 18:2534. doi.10.1111/j.1654.1103.2007.tb02513.x.

DeLuca, T. H., W. A. Patterson IV, W. A. Freimund, and D. N. Cole. 1998. Influence of llamas, horses and hikers on soil erosion from established recreation trails in western Montana, USA. Environmental Management 22:255-262.

Distel, R. A. 2013. Manejo del pastoreo en pastizales de zonas aridas y semiaridas. Revista Argentina de Produccion Animal 33:53-64.

Elzinga, C. L., D. W. Salzer, and J. W. Willoughby. 1998. Measuring and monitoring plant populations. Bureau of Land Management, Department of Interior Technical References 1730-1.

Estrada-Castillon E., J. A. Villarreal-Quintanilla, E. Jurado-Ybarra, C. Cantu-Ayala., M. A. Garcia-Aranda, J. Sanchez-Salas., J. Jimenez-Perez, and M. Pando-Moreno. 2011. Clasificacion, estructura y diversidad del matorral sub-montano adyacente a la planicie costera del Golfo Norte en el Noreste de Mexico. Botanical Sciences 90:37-52.

Fleischner, T. L. 1994. Ecological costs of livestock grazing in western North America. Conservation Biology 8:629-644.

Fuhlendorf, S. D., D. E. Townsend II, R. D. Elmore, and D. M. ENGLE. 2010. Pyric-herbivory to promote rangeland heterogeneity: evidence from small mammal communities. RangelandEcology & Management 63:670-678.

Gillen, R. L., F. T. McCollum, M. E. Hodges, J. E. Brummer, and K.W. Tate. 1991. Plant community responses to short duration grazing in tallgrass prairie. Journal of Range Management 44:124-128.

Gutierrez-Cedillo, J. G., L. I. Aguilera-Gomez, and C. E. Gonzalez-Esquivel. 2008. Agro-ecology and sustainability. Convergencia 46:35-71.

Hernandez, V. G., L. R. Sanchez V., T. F. Carmona V. M. R. Pineda L., and R. Cuevas G. 2000. Efecto de la ganaderia sobre la regeneracion arborea de los bosques de la Sierra de Manantlan. Madera y Bosque 6:13-28.

Higgins, S. I., J. Kantelhardt, S. Scheiter, and J. Boerner. 2007. Sustainable management of extensively managed savanna rangelands. Ecological Economics 62:102-114.

Hortal, J., P. A. V. Borges, C. Gaspar. 2006. Evaluating the performance of species richness estimators: sensitivity to sample grain size. Journal of Animal Ecology 75:274-287.

Kimball, T. L. 1980. Noncompetitive rangeland management for wild and domestic animals. Rangelands 2:24-25.

Lauenroth, W. K., D. G. Milchunas, J. L. Dodd, R. H. Hart, R. K Heitschmidt, and L. R. Rittenhouse. 1999. Effects of grazing on ecosystems of the Great Plains. Pages 69-100 in Ecological implications of livestock herbivory in the west (M. Vavra, W. A. Laycock, and R. D. Pieper, editors). Second edition. Society for Range Management, Denver, Colorado.

Magurran, A. E. 2004. Measuring biological diversity. Blackwell Science Ltd., Oxford, United Kingdom.

Manzano, M. G., J. Navar, M. Pando-Moreno, and A. Martinez. 2000. overgrazing and desertification in northern Mexico: highlights on north-eastern region. Annals of Arid zone 39:285-304.

McKinney, B. R., and J. Delgadillo Villalobos. 2014. Overview of El Carmen Project, Maderas del Carmen, Coahuila, Mexico. Pages 14-17 in Proceedings of the Sixth Symposium on the

Natural Resources of the Chihuahuan Desert Region (C. A. Hoyt and J. Karges, editors). The Chihuahuan Desert Research Institute, Fort Davis, Texas.

Menard, C., P. Duncan, G. Flaurance, J. Y. Georges, and M. Lila. 2002. Comparative foraging and nutrition of horses and cattle in European wetlands. Journal of Applied Ecology 39:120-133.

Milchunas, D. G., and W. K. Lauenroth. 1993. Quantitative effects of grazing on vegetation and soils over a global range of environments. Ecological Monographs 63:327-366.

Nabity, P. D., J. A. Zavala, and E. H. DeLucia. 2009. Indirect suppression of photosynthesis on individual leaves by arthropod herbivory. Annals of Botany 103:655-663.

Noss, R. F. 1994. Cows and conservation biology. Conservation Biology 8:613-616.

Olff, H., and M. E. Ritchie. 1998. Effects of herbivores on grassland plant diversity. Trends in Ecology and Evolution 13:261-265.

Peel, M. C., B. L. Finlayson, and T. A. McMahon. 2007. Updated world map of Koppen-Geiger climate classification. Hydrology and Earth System Sciences 11:1633-1644.

Powell, A. M. 1997. Trees and shrubs of the Trans-Pecos and adjacent areas. First edition. The University of Texas Press, Austin, Texas.

Ralphs, M. R., M. M. Kothmann, and C. A. Taylor. 1990. Vegetation response to increased stocking rate in short-duration grazing. Journal of Range Management 43:104-108.

Reiner, R. J., and P. J. Urness. 1982. Effect of grazing horses managed as manipulators of big game winter game. Journal of Range Management 35:567-571.

Renne, I. J., and B. F. Tracy. 2007. Disturbance persistence in managed grasslands: shifts in aboveground community structure and the weed seed bank. Plant Ecology 190:71-80.

Seagle, S. W., and S. Liang. 2001. Application of a forest gap model for prediction of browsing effects on riparian forest succession. Ecological Modelling 144:213-229.

Su, H., W. Liu, H. Xu, Z. Wang, H. Zhang, H. Hu, and Y. Li. 2015. Long-term livestock exclusion facilitates native woody plant encroachment in a sandy semiarid rangeland. Ecology and Evolution 5:2445-2456.

Todd, S.W. 2006. Gradients in vegetation cover, structure and species richness of Nama-Karoo shrublands in relation to distance from livestock watering points. Journal of Applied Ecology 43:293-304. doi:10.1111/j.1365-2664.2006.01154.x.

Valone, T.J. 2003. Examination of interaction effects of multiple disturbances on an arid plant community. Southwestern Naturalist 48:481-490.

Vermeire, L. T., J. L. Crowder, and D. B. Wester. 2014. Semiarid rangeland is resilient to summer fire and postfire grazing utilization. Rangeland Ecology & Management 67:52-60.

Vilanova, I., A. R. Prieto, and S. Stutz. 2006. Historia de la vegetacion en relacion con la evolucion geomorfologica de las llanuras costeras del este de la provincia de Buenos Aires durante el Holoceno. Ameghiniana 43:147-159.

Villarreal, H., M. Alvarez, S. Cordoba, F. Escobar, G. Fagua, F. Gast, H. Mendoza, M. Ospina, and A. M. Umana. 2006. Manual de metodos para el desarrollo de inventarios de biodiversidad. Programa de Inventarios de Biodiversidad. Instituto de Investigation de Recursos Biologicos Alexander von Humboldt, Bogota, Colombia.

Villarreal-Quintanilla, J. A., and J. Valdes-Reyna. 1993. Vegetacion de Coahuila, Mexico. Revista de la Sociedad Mexicana de Manejo de Pastizales 6:9-18.

Vogel, A., M. Scherer-Lorenzen, and A. Weigelt. 2012. Grassland resistance and resilience after drought depends on management intensity and species richness. PLoS oNE 7: e36992. doi:10.1371/journal.pone.0036992.

Yancey, M. J., and C. L. Douglas. 1983. Burro-small vertebrate interactions in Death Valley National Monument, California. Desert Bighorn Council Transactions 23:17-24.

Zhao, G., J. Liu, W. Kuang, Z. Ouyang, and Z. Xie. 2015. Disturbance impacts of land use change on biodiversity conservation priority areas across China: 1990-2010. Journal of Geographical Sciences 25:515-529.

Submitted 26 October 2016. Accepted 12 May 2017.

Associate Editor was James Moore.

Jose Javier Ochoa Espinoza,* Cesar Cantu Ayala, Eduardo Estrada Castillon, Fernando Gonzalez Saldivar, Jose Uvalle Sauceda, Enrique Jurado, Leonardo ChapaVargas, Edmar Melendez Jaramillo, and Edgardo Ortiz Hernandez

Facultad de Ciencias Forestales-Universidad Autonoma de Nuevo Leon, Km 145 Carretera Nacional 85, Apartado Postal 41, C.P. 67700, Linares, Nuevo Leon, Mexico (JJOE, CCA, EEC, FGS, JUS, EJ, EMJ, EOH)

Division de Ciencias Ambientales, Instituto Potosino de Investigacion Cientfica y Tecnologica AC. Camino a la Presa San Jose 2055, Lomas 4a Seccion, C.P. 78216, San Luis Potose, Mexico (LCV)

* Correspondent:

Caption: Fig. 1--General location of the study area and distribution of the two types of scrublands.

Caption: Fig. 2--Differences in average importance value index (IVI) between herbaceous and scrub strata of microphyllous and rosetophyllous scrublands with and without effect of livestock in northern Coahuila.
TABLE 1--Richness, abundance, and diversity of two desert scrubland
communities in two grazing conditions, and during two seasons.
Different letters within rows were different (P < 0.05).


Ecological         With        Without      Dry       Wet
parameters (a)   livestock    livestock    season    season    Total

Sobs                63 ab      45 ab        60 ab     53 ab      66
N                1,027 a      351 b        766 ab    613 ab   1 378
Chao 1             101         66           85.5     104         74.62
Jack 1              83.3       65.9         89.45     75.23      81.73
[R.sup.2]            0.99       0.99         0.99      0.99       0.99
Sest                87.2       77.9        102.7      79.7       82.5
Slope                0.59       0.64         0.84      0.60       0.23
[]           0.89       0.90         0.88      0.90       0.88


Ecological         With        Without       Dry        Wet
parameters (a)   livestock    livestock    season     season

Sobs                68 a       56 b           62 ab      64 ab
N                1,925 a      997 b        1,374 ab   1,545 ab
Chao 1              87.43      71.17          73.14      72.27
Jack 1              88.3       73.4           80.37      82.37
[R.sup.2]            0.99       1.00           1.00       1.00
Sest                 81.5      73.3           82.7       79.4
Slope                 0.41      0.45           0.53       0.44
[]            0.86      0.86           0.85       0.87

parameters (a)   Total

Sobs                75
N                2,919
Chao 1              86.38
Jack 1              91.72
[R.sup.2]            0.99
Sest                86.7
Slope                0.19
[]           0.86

(a) Sobs = observed richness; N = abundance; [R.sup.2] = determination
coefficient; Sest = estimated richness;
[D.sub.Si] = Simpson diversity index.

Table 2--Effect of livestock presence on importance value index of
10 plant species in microphyllous and rosetophyllous scrubland.
Plant species sharing the same letter are not significantly


Plant species          With livestock   Without livestock

Acacia farnesiana      24.75 cdefg      15.94 abcdefg
Agave lechuguilla      32.82 ei         35.06 efghi
Bouteloua eriopoda     6.24 a           10.17 ab
Bouteloua gracilis     12.11 cdefg      34.19 efghi
Jatropha dioica        21.52 bcd        13.17 abcd
Larrea tridentata      50.08 h          34.98 efghi
Opuntia leptocaulis    29.55 defg       18.36 abcdefg
Opuntia macrocentra    23.16 cdefg      24.49 cdefg
Parthenium incanum     22.59 bcdefg     16.77 abcdefg
Prosopis glandulosa    71.40 hi         18.07 abcdefg


Plant species          With livestock   Without livestock

Acacia farnesiana      17.15 bc         15.09 bc
Agave lechuguilla      46.54 g          39.19 fg
Bouteloua eriopoda     17.02 bc         12.69 ab
Bouteloua gracilis     19.78 bc         28.26 cde
Jatropha dioica        15.52 bc         16.09 bc
Larrea tridentata      33.06 ef         17.74 bc
Opuntia leptocaulis    12.31 abc        4.41 a
Opuntia macrocentra    14.54 bc         18.79 bcd
Parthenium incanum     20.78 cd         28.11 de
Prosopis glandulosa    15.85 abcd       10.35 abcd
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Author:Ochoa Espinoza, Jose Javier; Cantu Ayala, Cesar; Estrada Castillon, Eduardo; Gonzalez Saldivar, Fern
Publication:Southwestern Naturalist
Geographic Code:1MEX
Date:Jun 1, 2017

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