The study was conducted in the Borana rangelands of southern Ethiopia (Figure 1), covering a total area of 95,000 km², bordering Somalia and Kenya (Megersa et al., 2014; Tuffa and Treydte, 2017). The Borana rangelands are situated within the altitudinal range of 750–2000 m above sea level (m.a.s.l) (Dalle et al., 2006b) and experience an arid and semi-arid climate with pockets of sub-humid zones (Megersa et al., 2014). The inter-annual rainfall varies, with an average annual rainfall of 527 mm (Dalle et al., 2006a) and a high coefficient of variability (18%–69%) (Angassa and Oba, 2008). The rainfall in Borana is bimodal, with a long rainy season between March and May and a short rainy season between September and November. The average annual temperature of the area varies from 19°C to 26 °C. Drought is common in the area, with isolated dry years (<400 mm) happening once every 5 years but becoming more frequent (Tuffa and Treydte, 2017). The region is characterised by limited availability of surface water; the major sources of water for humans and livestock are deep wells and ponds (Homann, 2004). The soils of the study area are derived from ancient alluvial and volcanic materials (FAO, 1986), with shallow red sandy loam in the uplands and vertisols in the bottomlands (Dalle et al., 2006b). Vertisols are characterised by their black colour, relatively high in soil nutrients compared to other locally abundant soil types such as the friable sandy loam ‘red soils’ (Augustine and Mcnaughton, 2004), with high clay content, reduces the infiltration of water and expands and contracts with changes in moisture content disrupts root systems (Pringle et al., 2016). While the high nutrient content and moisture availability can favour the establishment of certain encroching woody plants, like Vachellia drepanolobium (Kenfack et al., 2021), poor drainage can limit others species prefer well-drained soils. Tthe impacts of vertisols on encroaching plants depends on the specific plant species and its adaptations.

The livelihood of the Borana pastoralists mainly depends on livestock production in this tropical savanna with varying proportions of open grassland and perennial herbaceous and woody vegetation, composed of Vachellia-Commiphora small-leaved deciduous woodlands (Coppock et al., 2007). Encroachment of woody plants has become a major challenge, causing a decline in herbaceous species, rangeland productivity, range condition, and livestock production (Tefera et al., 2007; Yusuf et al., 2011). Borana pastoralists have historically used controlled fire to manage rangelands for suppressing undesirable woody plants, promoting grass growth, and reducing tick population (Angassa and Oba, 2008). However, in the early 1970s local burning practices were prohibited (Angassa and Oba, 2008; Feyisa et al., 2023), which largely contributed to the prolifirication of certain woody species including Vachellia brevispica, V. bussei, V. drepanolobium, V. etbaica, V. mellifera, V. nilotica, V. seyal, and Commiphora africana (Yusuf et al., 2015; Feyisa et al., 2017). Fire can have a considerable impact on seedlings, growth, survival and adult recruitment of woody plants (O’Connor et al., 2014). The expansion of woody plant species due to the ban on fire use is likely attributed to combination of factors, including overgrazing, rainfall and soil fertility influences (Liao et al., 2018). Vachella drepanolobium is a perennial, leguminous, thorny plant that reaches up to 7 m in height and is mostly found on black cotton soils, where it grows in mono-dominant stands and makes up 98% of the individual tree species (Hare et al., 2020). This species is highly adapted to fire, with relatively thick bark (Midgley et al., 2016) and strong post-fire resprouting (Okello, 2007). It is the most widespread myrmecophyte tree in East Africa, producing domatia (swollen thorns) and extrafloral nectaries that provide shelter and food rewards to ants (Palmer et al., 2008). Vachellia. drepanolobium trees host four symbiotic ant species; these include Crematogaster sjostedti, Crematogaster mimosae, Crematogaster nigriceps, and Tetraponera penzigi, which provide varying levels of protection against browsing insects and mammalian herbivores including elephants (Palmer et al., 2008). These ants can increase the survival of trees and stabilise tree cover across the landscape (Goheen and Palmer, 2010). Vachellia drepanolobium is aggressively encroaching into rangelands, suppressing grass production and affecting the productivity of the rangelands (Tefera and Mlambo, 2010). This is most likely due to a combination of factors, including grass-tree competition (Riginos et al., 2009), being protected from browsing herbvior by ants and spines (Okello, 2007; Kenfack et al., 2021), the lack of fire, the exclusion of browers, and soil properties (black vertisol soil, Okello, 2007). As Riginos et al. (2009) reported that a stronger suppressive effect of grass competition on tree growth. The elimination of grass competition resulted in a doubling of the tree growth rate across all tree age classes. Vachellia drepanolobium has a low forage value because it is heavily protected against herbivores by spines and ants (Okello, 2007). but it provides the main sources of feed for black rhinos (Diceros bicornis) and giraffes (Giraffa camelopardalis) (Birkett, 2002). Thus, all woody plants should not be removed from savanna landscapes, woody plants play important ecological roles (Smit, 2005). Complete removal rather than selective thinning of trees could negatively affect rangeland productivity, biodiversity, and ecosystem services, particularly in the long term (Smit, 2014). Maintaining ceratin trees species on the rangelands can have multiple benefits to pastoralists, ranchers, and wildlife (Monegi et al., 2023a).

The study was conducted in the year 2020 over two growing seasons in the Borana rangelands of the Yabello district, Dida Hara Kebele Administration (KA), which has a total area of 985 km² and is located between 1,200 and 1,600 m. a.s.l. (Angassa and Oba, 2010; Figure 1). The experimental sites were selected within the community enclosure named Kelo Olla Doyo Dube. The enclosure, established before 25 years ago, and cover over 4.3 square kilometers, which was fenced and grazed only during the dry season, providing some fuel load for fire (Angassa and Oba, 2010). We selected two representative grazing areas through participatory mapping together with local elders, who knew the history and nature of the rangeland as experimental sites. The sites had similar soil types (vertisol), topography, and no history of earlier disturbances such as bush clearing and burning of the grazing areas. The vegetation consisted mainly of savanna, dominated by V. drepanolobium trees and herbaceous species. The human population settlements in the Dida-Hara area are semi-permanent settlements with an average livestock holding per household of 13 cattle, 11 small ruminants (goats and sheep), and 2 camels (Tefera et al., 2007). The majority of the open communal rangelands have deteriorated as a result of the high stocking rate, which was estimated to be 0.235 Tropical Livestock Units ha⁻¹ (1 TLU 250 kg; Homann et al., 2008). Despite the unregulated stocking rate outside on the communal grazing land, overgrazing was not a threat within the enclosure, and grazing effects after thinning could be excluded as a factor in our analyses (Angassa and Oba, 2010). Overgrazing outside the study plot may have spatially compressing effects on soil condition (e.g., soil moisture) and vegetation composition within the plot, potentially affecting tree-grass competition. This external grazing could be limiting factor on the observed results.

We selected two 1.42 ha experimental sites that had mono-specific stands of “whistling thorn” stands (V. drepanolobium). At each site, we established 12 plots of 10 m × 20 m (200 m²), with 3 m-wide firebreaks between the plots and a 5 m-wide distance between replication treatments. Similar to a study conducted in Ethiopia (Hare et al., 2021a) and in Burkina Faso (Savadogo et al. (2008), we thinned each experimental plot by 67% of the initial tree density (2,808 ± 161 tree per ha⁻¹) found in the control plot (Abate et al., in review). We applied four different post-thinning management treatments to each plot: 1) Browsing, i.e., allowing goats to browse freshly growing stump buds after tree cutting within the entire plot. 2) Fire = burning of the entire plot once following tree cutting. 3) Fire and browsing = the entire plot was burned with prescribed fire, and goats were allowed to browse the entire plot afterwards, and 4) Control = 33% tree cover remaining, and tree stumps in the plot were left untreated. The treatments were replicated three times and assigned randomly to the plots. Before the post-tree thinning operation, the density of trees and the number of seedlings within each plot were counted and measured. During thinning, we marked and tagged each cut tree with aluminium tags at the base of the trunk, below the cutting point. Using a local axe, trees had been cut at the base (0.5 m above ground), and the stumps had been debarked to facilitae death of the tree at the end of the dry season in February 2021. While cutting, we did not favour any particular tree sizes and made sure the remaining trees were scattered to reduce competition (Hare et al., 2020; Abate et al., in review). The cut woody materials were chopped up into finer pieces and then piled on tree stumps as fuel load for burning (Angassa and Oba, 2009). The burning was conducted late in February, the peak of the long dry season. Before burning, we assessed the herbaceous biomass to determine the amount of fuel load for a plot. In average the fuel load in fire alone treatment (2) was 2,623 ± 118 kg ha⁻¹ with moisture content of 14%, while the amount of fuel load in fire and browsing plot (3) was 2,803 ± .290 kg ha⁻¹ with moisture content of 19%. A fuel load of more than 2000 kg DM ha⁻¹ has been reported sufficient to ignite fire and killing tree (Angassa and Oba, 2009; Kahumba, 2010; Hare et al., 2021b). We conducted the burning between 15:00 and 17:00 local time, considering the local weather conditions (e.g., air temperature, relative humidity, wind direction and speed, fuel load availability, and fire risk control). We did not record the temperature at the time of burning due to the absence of infrastructure and technology.

For the browsing treatment plots, 10 mature male Borana goat individuals were allowed to browse each plot (50 heads of goats ha⁻¹) (Bonanno et al., 2007). We started browsing following the main rains, when tree stump buds emerged after coppicing in the control plots. The browsing on each plot was conducted biweekly for 1 hour (equivalent to one goat utilising a plot for 10 h) in each of the two growing periods in 2020.

The experimental plots were fenced off and protected to prevent disturbance from humans and livestock grazing and browsing. During the post-thinning management periods, the number of V. drepanolobium trees, saplings, and seedlings were counted and measured at the end of each growing period. We monitored the V. drepanolobium tree and shrub stump mortality (CutMort) of all trees per plot that had been cut in the previous bush-clearing treatment. We further measured tree and shrub mortality (TreeMort) of trees that had not been cut at the previous treatment, coppicing (TreeCopp) (referring to regrowth in response to treatment effects), seedling mortality (SeedMort), and new seedling recruitment (SeedRec), i.e., V. drepanolobium plants of <0.5 m in height (Angassa and Oba, 2010).

We evaluated rangeland condition in each plot based on grass and soil factors used for range condition scores adopted by Baars et al. (1997) and other studies (Dalle et al., 2006b; Angassa, 2012; Melak et al., 2019). The criteria include botanical composition of grass species, basal cover, litter cover, number of grass seedlings, age distribution, and soil factors (soil erosion and compaction) (Baars et al., 1997; Supplementary Table S1). The maximum possible score for a plot was 50 points, indicating excellent (41–50 points), good (31–40 points), fair (21–30 points), poor (11–20 points), or very poor rangeland condition (<10 points) (Baars et al., 1997; Supplementary Table S1). The methodologies have described below and presented in Supplementary Table S1.

The grass species composition was assessed in three 1 × 1 m quadrants within each plot. Individual grass species were counted and identified during both the main (May) and short rainy season (November), i.e., growing period 1 and 2, respectively. The identified grass species were classified according to life forms (annual and perennial) and ecological status into (i) decreasers (desirable species likely to decrease with heavy grazing pressure), (ii) increasers (intermediately desired species likely to increase under heavy grazing intensity), and (iii) invaders (undesirable species likely to increase under heavy grazing intensity) based on the histories of particular grass species (Jenkins et al., 1974; Tainton, 1981) and the opinions of local communities. For grass species composition rating, a one-to-ten-point scale was given (see Supplementary Table S1).

Basal cover and litter cover of grass species were evaluated on a scale of 0–10 points (Supplementary Table S1). Basal cover refers to a ground cover by the base of the grasses. Grass litter was defined as dead plant material that had fallen to the ground (Mengistu, 2005). In each experimental plot, a representative sample area of three 1 m² quadrants were established for the detailed assessment of both basal and litter cover (Baars et al., 1997), which was visually estimated and then harvested. The rating for litter cover within the same square meter was given the maximum score (10 points) when grass litter cover exceeded 40%, and the minimum score was given when the litter cover was less than 3% (Supplementary Table S1).

For both the number of grass seedlings and the age distribution of grasses, a score of 1–5 points was used (Supplementary Table S1). The number of grass seedlings was counted in three areas of each experimental plot following Baars et al. (1997). An A4 sheet of paper (30 × 21 cm) was dropped at random from a distance of 2 m height. A maximum score of five points was given when all age categories of grasses (i.e., young, medium, and old) were present (see Supplementary Table S1).

For the soil condition, a score of 0–5 points was used, estimated by the status of soil erosion considering pedestals and pavements Baars et al. (1997). Soil compaction (0–5 points) was assessed based on the amount of capping (crust formation), following the methods described by (Baars et al., 1997; Supplementary Table S1).

Dry matter (DM) biomass of the herbaceous species was determined in each treatment plot by harvesting biomass from three 1 m² quadrants at ground level. The harvested fresh material was weighed and hand-separated into grasses and forbs. The grasses were further separated into different species, and the proportional contribution of each species was determined. The samples of collected grasses and forbs were oven-dried at Hawassa University, College of Agriculture, in the Animal Nutrition Laboratory at 65°C for 24 h and weighed. Identification and nomenclature of plant species followed the guidelines of Hedberg (1996). Additional references such as the Guide to the Grasses of Ethiopia (Froman and Persson, 1974) and Grasses Common to Arero Area (Jenkins et al., 1974) were used for plant identification. A summery of collected variables is presented in Table 1.

A total of 24 sample units (4 treatments x 3 replicates x 2 sites) were considered for the data analysis on the response of woody trees to post-thinning management, while for the herbaceous layer, we included growing period as an additional factor (4 post thinning management treatments x three replications x two sites x two growing periods). Data were checked for normality, homogeneity of variance, and assumptions of linearity. For data that did not fulfill the normality assumption, we used square root transformation for tree mortality, log transformation for DM biomass and range condition parameters (Osborne, 2002). The woody plant parameters, herbaceous species composition (abundance of grass and forb species), diversity, range condition, bare ground cover, and DM biomass parameters were analyzed using the general linear model procedure of the Statistical Analysis System (SAS, 9.4) and SPSS, 26. To investigate the impact of post tree thinning treatment, we used the tree species attributes (e.g., TreeSBMort, TreeCopp, TreeMort, SeedMort, and SeedRec) as dependent variables and post-thinning managements as independent variables. The range condition assessment parameters, including grass composition, basal cover, litter cover, the age distribution of grasses, number of grass seedlings, soil condition, DM biomass parameters, and bare ground cover, were considered as dependent variables and analyzed using one way ANOVA to examine the effect of post-thinning management and two-way ANOVA to assess effect of post tree-thinning management and growing period.

Herbaceous species that were recorded in each of the treatment plots at the end of the second growing period were ordinated using Correspondence Analysis (CA) (Greenacre, and Blasius, 2006) in PAST 4.13 software (Hammer et al., 2001). Total herbaceous species, grass, and forb species alpha diversity, richness, and evenness were analyzed using the Shannon-Wiener index (Kent, 1992) using PAST 3 software. Tukey-HSD post hoc test was used to test for differences in averages of woody and herbaceous (grass and forb) layer responses, biomass and rangeland condition parameters at p < 0.05.

All the three post-tree thinning management scenarios significantly enhanced tree mortalities, reduced seedling recruitment, Prescribed fire (2) and fire and goat-browsing (3) treatments resulted in significantly greater (compared to goat browsing (1) and control treatment (4). We found that post-thinning management significantly affected tree mortality (Wilks’ λ = 04; F = 4.24; p < 0.001).

Relative to the control plot, tree mortality approximately doubled under the fire and browsing treatment and fire only plots and increased by about 20% under goat browsing only (Figure 2). For trees that had not received thinning management (i.e., not-cut), mortalities increased by >20% in the fire and browsing and fire only (F = 17.8; p < 0.001) plots and did not change in the goat browsing only plot (Figure 2). Tree coppicing was five times lower in the fire and browsing and 11-fold lower in the fire only plots relative to the control plot (F = 54.3; p < 0.001; Figure 2). Seedling recruitment was significantly reduced under post-thinning treatments, with control plots showing more than 14-times higher than other treatments (F = 4.6; p = 0.013; Figure 3). Surprisingly, seedling mortality did not show a significant difference among treatments (F = 2.0; p = 0.140; Figure 3) and was generally high with 62% ha⁻¹ across all treatments. Almost 78% of seedlings died in the fire and browsing plots; about 67% died in fire only plots while 75% died in the browsing only treatment and only 42% died in the control plots (Figure 3).

Post-thinning management significantly increased the abundance of the dominant desirable grass species. Grass and forb species richness, forb diversity. Prescribed fire (2) and fire and goat-browsing (3) treatments resulted in significantly higher as compared to goat browsing (1) and control treatment (4).

We identified a total of 42 herbaceous species, 21 grasses, and 21 forb species across treatments. (Table 2; Supplementary Figure S1). The grass species comprised 67% perennials and 33% annuals, and they were composed of 29% highly desirable, 12% intermediately desirable, and 48% less desirable grass species. The dominant grass species in our study were Chrysopogon aucheri, Pennisetum mezianum, Cenchrus ciliaris, Pennisetum stramineum, Sporobolus pyramidalis, Eragrostis papposa, Digitaria milanjiana, and Cynodon dactylon. The correspondence analysis (CA) results did not clearly associate the herbaceous species with different treatments, as about 93.1% of the total variance in species abundance was explained by the model (Supplementary Figure S1). We found that the abundance of C. ciliaris, C. aucheri, P. mezianum, and P. stramineum was more than twice as high in the fire and browsing treatments compared to the control and browsing only treatments (Table 2). Moreover, S. pramidalis and D. milanjiana were more than double that in the control compared to the browsing only treatment (Table 2). In the fire only plot, the abundance of C. dactylon was more than four times higher than in the browsing and in the control treatments (Table 2).

We further found that post-thinning management significantly influenced grass species richness (F = 3.5; p < 0.022), forb species richness (F = 5.2; p = 0.004), with grass and forb species richness increased by 18% and 40% in the control plot than in the browsing only treatment, respectively (Figure 4A), but no significant difference with fire only and fire and browsing treatments. Similarly, the diversity of forb species differed (F = 7.1; p < 0.001), being higher in the control plots (F = 9.4; p < 0.001) compared to the browsing only plots but similar to fire only and fire and browsing treatments (Figure 4B). The diversity of grass species did not show significant differences among treatments (F = 0.8; p = 0.551).

Post-thinning treatments significantly influenced parameters of range condition (Table 3), except soil condition (erosion and compaction), and interactions with the growing period were insignificant. Relative to the control plot, the grass species composition rating increased by 25% in the fire only treatment (Table 4). The total range condition scores were highest in the fire and browsing and fire, treatments (Table 4). The total range condition was 10% better under the fire treatment than in control, and the fire plot was rated as excellent condition while the other treatments showed good condition. Bare ground cover was more than 1.5 times as high in the browsing treatment than in the fire and browsing and fire only treatments (Table 4).

Grass DM biomass was more than double that in the fire treatment compared to browsing only and in the control treatments (F = 54.5; p < 0.001; Figure 5A), while forb DM biomass (F = 111.6; p < 0.001; Figure 5A) showed more than 1.6 times higher values in the control plot than the other treatments. The DM biomass of highly desirable (F = 64.7; p < 0.001) grasses was more than double under the fire only treatment compared to the other treatments (Figure 5B).

Growing period significantly affected overall grass biomass (F = 19.0; p = 0.001), biomass of highly desirable (F = 27.1; p = 0.001), and less desirable grass species (F = 28.8; p = 0.001) (Figure 6), grass species richness (F = 5.6; p = 0.014), and forb richness and diversity (F = 15.6; p = 0.001 and F = 4.3; p = 0.001, respectively) (Figure 7), with all those trends being higher in the second growing season than in the first (Figure 7). With the exception of low desirable grass biomass (F = 5.7; p = 0.002). The diversity of grass did not show significant differences between growing periods (F = 0.8; p = 0.393). There was no interaction between thinning intensity and growing period on biomass, range condition metrics, and diversity indices.

Our study showed that prescribed fire, combined fire and browsing and goat browsing treatments could effectively complement the benefits of tree thinning by increasing the encroacher species’ tree mortality, resulting in reduced tree cover. Tree mortality was particularly strong in treatments that involved prescribed fire. Our observed results are similar to earlier studies made on tree thinning, herbivory and fire disturbance in southern Ethiopia and Namibia (de Klerk, 2004; Hare et al., 2020; Hare et al., 2021a), Burkina Faso (Savadogo et al., 2008) and in Missouri, United States (Beebe et al., 2021a). The higher mortalities of mature trees and saplings under fire, and fire and browsing after thinning can impose separate or combined disturbances and weaken trees and seedling’ resilience to environmental conditions (e.g., drought, insects; Kane et al., 2017; Hood et al., 2018), amplifying mortalities. Although goat browsing did not show a significant difference from the control treatment, we showed that browsing alone could increase tree mortality by 20% but is likely insufficient on its own in completely controlling regrowth (Hester et al., 2006). We found that prescribed fire and combined fire and goat browsing significantly increased non-cut sapling mortalities and reduced tree coppicing compared to the control plots, which suggests that resprouting abilities are hampered when fire is included (LaMalfa et al., 2019; Monegi et al., 2023a).

As expected, all our treatments reduced seedling recruitment, which is consistent with other studies, that claim that browsing herbivores (Staver and Bond, 2014; Augustine et al., 2019) and a combination of fire and browsing (Staver et al., 2009) negatively affect seedling survival. Moreover, the higher herbaceous biomass we found after tree thinning (Abate et al., in review) can further be attributed to seedling mortality through strong grass competition (Riginos et al., 2009; Pillay and Ward, 2021; Boys, 2022; Donaldson et al., 2022; Monegi et al., 2022), As reported by Riginos et al. (2009) grass competition has strong suppressive effect on seedling and tree growth. when grass competition was removed, it resulted in a doubling of the growth rate across all demographic stage of the tree classes, suggesting that fire and herbivory as well as grass competition may create demographic bottlenecks that influence tree populations (Donaldson et al., 2022) and may in the future contribute to a reduction in woody cover and help keep savanna systems open up (Barton and Hanley, 2013). We claim that in Ethiopia, the re-introduction of fire as a management tool can likely maintain the savanna ecosystems of Borana.

Our result showed that the use of prescribed fire, goat browsing, and combinations thereof significantly enhanced the benefit of tree thinning by increasing the abundance of certain grass species, similar to research in South Ethiopia (Hare et al., 2021b). The increased abundance of certain grass species under the conditions of a fire-only plot and fire and browsing after thinning can lead to separate or combined disturbances. These disruptions can decrease the canopy gap, reduce competition for resources, and increase irradiation, nutrients, and soil moisture. This creates a favourable condition for the establishment and growth of herbaceous plants (grasses and forbs), resulting in an increase in the abundance and diversity of plants (Hare et al., 2021b; Monegi et al., 2022). Other studies indicated that herbaceous plant abundance increased with fire treatments (Bielski et al., 2021; Gold et al., 2023) and fire-goat browsing treatments (Beebe, 2021b). The high abundance of C. ciliaris, C. aucheri, P. mezianum, P. stramineum, S. pramidalis, C. dactylon, and T. triandra in fire and fire and browsing plots suggests that these species are fire-tolerant, which was found for some of those species already in South Africa (Ratsele, 2013). In contrast goat browsing and control plots favoured the abundance of Aristida adoensis, Bothriochloa insculpta, Chloris roxburghiana, Digitaria milanjiana, Eragrostis papposa, Heteropogon contortus, Panicum maximum and Setaria verticillata grasses, complementing findings that P maximum (Fynn et al., 2005; Smith et al., 2013; Monegi et al., 2023a) and B. insculpta (Monegi et al., 2023b) increased in abundance in unburned plots.

We also found that the abundance of certain forb species such as Abutilon hirtum, Commelina africana, Pavonia erlangeri, and Tagetes minuta was substantially reduced under post-thinning treatments, which might be important for managers as forbs are often less preferred by grazer livestock (Oba et al., 2000). We found that the goat browsing treatment significantly reduced grass and forb richness. Selective and high browsing pressure in the goat-browsing treatment could explain the reduced herbaceous species richness (Salgado-Luarte et al., 2019; Nolden, 2020). Grass diversity was not affected by our post-thinning treatments, which agrees with Monegi et al. (2023a), but it might also indicate that a longer experimental period could have shown more pronounced shifts in the diversity of grass species.

As we expected our treatments increased the rangeland condition and biomass of the herbaceous layer, similar to Hare et al. (2021b). The improved rangeland condition and biomass of the herbaceous layer under fire only and fire and browsing after thinning can demonstrate the combined impacts of disturbances with the reduction of competition due to competitive release from existing trees (shading, competition for moisture, and nutrients). This creates a conducive environment for the growth of herbaceous plants. This led to an increase in basal cover of grasses, which in turn reduced bare ground cover, thereby enhancing biomass and range conditions (Monegi et al., 2022), which supporting livestock grazing and contributing to biodiversity conservation (LaMalfa et al., 2019; Monegi et al., 2022), Similar to Hare et al. (2021b), we found a stronger response in rangeland biomass, diversity and species compostion to variation in climatic conditions (rainfall and seasonality) and soil moisture contents.