Over the past quarter century, a substantial body of research has tended to focus on the negative effects of shrub encroachment on ecosystem properties and processes as broad as suppression of understorey plant cover (Archer, Boutton & Hibbard 2001; Huenneke et al. 2002), reduced soil nutrient pools (Sparrow et al. 2003), altered nutrient distributions (Schlesinger, Tartowski & Schmidt 2006) and greater wind and water erosion (Abrahams, Parsons & Wainwright 1994). In direct contrast, we detected a positive effect of increasing shrub cover on average structure and function, and substantial increases in specific measures of structure and function indicative of healthy landscapes, with increasing shrub cover. This was particularly true under low grazing pressure although it varied between the two sites. Our SEM models of ecosystem behaviour indicated that increasing shrub cover had positive effects on most ecosystem properties (infiltration, stability and nutrient indices, infiltrability) whereas grazing had either negative or benign effects (Figs 1 and 2). Although the commonly held view is that encroachment suppresses herbaceous biomass (e.g. Knapp et al. 2008), we showed that this effect is not universal, with biomass declining at Scotia (path coefficient = −0·57), but remaining unchanged at Buronga (Fig. 2). We do not believe that this is due to rainfall, given the similar rainfall regime at both sites, but rather, to the many factors, and their inter-relationships that drive plant–plant interactions in drylands (see Callaway 2007). The results are consistent with recent research that questions the widely held view that encroachment is indicative of degraded systems (Maestre et al. 2009; Eldridge et al. 2011), providing strong evidence that woody plants, including shrubs, have positive, or neutral effects on the structure of dryland ecosystems and their biota. Our shrub species (E. sturtii, S. artemesioides, D. viscosa) are all medium-sized plants (up to 2 m tall) typical of woodland communities under varying levels of encroachment. Our conclusions, therefore, are applicable to extensive areas of Australia where these shrubland communities predominate.
Grazing dampens positive effects and amplifies negative effects of shrub encroachment
Higher levels of grazing in our study dampened the generally positive effects of shrubs on ecosystem functions, though the effect was site-specific, and mostly restricted to the Buronga sites (Fig. 5). We see two explanations for the differential grazing effect between sites. First, Scotia has had a shorter history of grazing (~70 years) than Buronga (>150 years), mainly because the Scotia sites were settled later and large areas were not watered before the middle of the 20th Century. Secondly, the surface soils at the Scotia sites were sandier than those at Buronga, and the different effects could have been due to differences in soil hydraulic behaviour and water holding capacity, which are known to vary markedly with soil texture (Noy-Meir 1973). Also, physical and biological soil crusts are less evident on sandy soils, therefore the effect of trampling by herbivores would be lower. In most cases, the effect of grazing was to negate the generally increasing effect on most functional response variables, so that under grazing, there was little or no effect of increasing shrub cover (Fig. 3). A reduction in the positive effects of shrubs under high grazing levels could be due to a breakdown in the facilitatory effect of shrubs, or may relate to a change from many smaller, to fewer larger, shrubs, which would permit a greater access by herbivores (Smit et al. 2007).
The greatest grazing-induced changes in function were evident in our three soil functional indices, which, although highly variable at high grazing levels, showed values under low levels of grazing. The pronounced dampening effects of grazing on the positive shrub cover functional relationships suggests that the disparity in widely diverging reports of shrub and shrubland effects on ecosystem properties and processes could be explained, in part, by grazing.
Published accounts of the relative effects of grazing and increasing woody plant density are few, though there is some support for our finding of a moderating effect of grazing on function. Studies by Throop & Archer (2008), and results from an area closer to the studied areas (S. Daryanto, unpublished data) showed that heavily grazed sites had less soil organic carbon (SOC) under woody plant canopies than lightly and heavily grazed sites, supporting the notion that the strong positive effect of woody plant encroachment on SOC accumulation was partially offset by grazing. Enhanced soil infiltration, stability and fertility under woody canopies compared with open sites has been widely documented (the so-called ‘islands of fertility’ phenomenon), and the plant effect is known to extend far beyond the edge of woody canopies (Wu & Archer 1985; Maestre et al. 2009). Interception of rainfall by relatively dense canopies of E. sturtii and D. viscosa would be expected to reduce soil surface sealing beneath shrub canopies. Furthermore, sub-canopy litter is known to moderate raindrop energy, reducing the tendency of the soil to form a physical crust (Geedes & Dunkerley 1999) and enhancing water penetration and organic matter content, which is also enhanced by the retention of run-off by shrub canopies. This higher soil fertility, together with the shading produced by shrub canopies are often invoked as the main mechanisms driving positive interactions among shrubs and their understorey species (Pugnaire, Haase & Puigdefábregas 1996). Improved microclimatic conditions may allow the recruitment of species that are less adapted to stressful conditions (Prider & Facelli 2004), and this effect is known to extend to biocrusts, where more water-demanding species grow prominently underneath plant canopies, but not in the open (Maestre 2003). The presence of these more productive species, together with the tendency for some plants to be protected from grazing and trampling under the canopies of dense shrubs, also explains the higher biocrust cover found under higher shrub covers.
A diminution of the well-known positive effects provided by shrubs under grazing might be due to several mechanisms. First, the scattering of litter into the interspaces by livestock, and exacerbated by wind or water can reduce the concentration of organic matter beneath shrub canopies (Li et al. 2009). Secondly, some of the species in our study (e.g. D. viscosa) are readily browsed by feral goats (Harrington & Johns 1990). Browsing reduces their capacity to intercept rainfall (Mills et al. 2009), and may affect their ability to retain litter (Bochet, Rubio & Poesen 1998). The differential effect of grazing on function could operate by altering the physical environment of the shrubs, by changing incoming levels of solar radiation (Yager & Smeins 1999) or by changing litter cover. Higher amounts of litter on the soil surface are expected to reduce the variability in soil temperature, which may in turn decrease soil organic matter decomposition rates and incorporation of carbon into deeper soil layers. Grazing therefore has substantial impacts on soil surfaces, altering their distribution, and ultimately affecting the infiltration of water in the interspaces beyond the shrub canopies. Grazing did, however, have a generally positive effect on infiltrability, which we attribute to the disruption to the biocrust, resulting in increases in infiltration (Eldridge et al. 2010). Thirdly, extremely high grazing levels lead to a weakening of positive plant–plant interactions (e.g. Smit et al. 2007), reducing the positive effect of shrubs on plants. Lastly, disturbance by some herbivores such as rabbits tend to be greater beneath shrub canopies; thus, the higher abundance of these grazers could overwhelm the positive effect of shrubs on biocrusts.
The management and policy of shrub encroachment
In Australia's semi-arid woodlands, debate over the effects of shrub encroachment on ecosystems is highly polarized. Those whose livelihoods depend on livestock grazing argue that shrub encroachment threatens the viability of their pastoral enterprises, given the heavy reliance of grass production. Indeed, sections of the agri-pastoral lobby argue that encroachment is synonymous with land degradation, despite the conclusions of recent global review (Eldridge et al. 2011) and reports to government (Eldridge, Wilson & Oliver 2003), which have failed to identify any clear and unambiguous links between encroaching vegetation and degradation. Despite the existence of contrasting evidence, implicit in current NSW (Australia) government legislation (Native Vegetation Act 2003) is the premise that the removal or reduction in encroaching woody plants (syn. ‘Invasive Native Scrub’; INS) is warranted to ‘improve or maintain’ environmental outcomes, a test under the NSW Native Vegetation Act (2003) http://www.legislation.nsw.gov.au. The listing of ‘feral native plant species’ as ‘invasive’ under NSW Legislation allows landholders to remove these plants under permit. The process for inclusion of species on the INS register has largely been based on their putative adverse effects on pastoral production. A similar situation exists in the western Cape of South Africa where the native shrub bankrupt bush Seriphium plumosum, has been listed as a Proclaimed Encroacher Plant (Regulation 16 of the Conservation of Agricultural Resources Act 43) because it displaces grasses and reduces the grassland grazing capacity (Wepener, Kellner & Jordaan 2008; Snyman 2012). In the United States, ranchers may apply for cost sharing from federal, state and local agencies to conduct brush removal on rangelands, with the stated purpose of enhancement of forage or water resources (Torell et al. 2005). An example is Texas SB1083 (Title 7 of the Agricultural Code), which establishes a fund to provide up to 70% cost sharing for brush removal, explicitly to promote groundwater recharge. Likewise, in New Mexico, substantial federal funding is being provided by the Bureau of Land Management to aerially apply herbicides to extensive areas of former grassland invaded by creosote bush (Restore New Mexico Program; http://www.blm.gov/nm/st/en/prog/restore_new_mexico.html).
An important issue identified in recent shrub encroachment models (Maestre et al. 2009; Eldridge et al. 2011) is the value that human society places on the ecosystem services provided by woody plants. Shrub encroachment can provide numerous conservation and societal benefits such as increased water recharge, habitat for organisms, carbon storage (Barger et al. 2011) and even autogenic regeneration at almost no cost to the community (Geddes et al. 2011). The extent of these benefits, however, has only recently been explored. For example, analyses of encroachment scenarios globally indicate that it is consistently associated with increases in the above- and below-ground carbon (Barger et al. 2011; Eldridge et al. 2011), and removal of shrubs may have little effect on recharge (Bazan et al. 2012). In Australia, the recent introduction of a price on carbon has provided renewed financial incentives for landholders to be involved in long-term carbon sequestration programmes such as the Carbon Farming Initiative (http://www.climatechange.gov.au/cfi) which allows farmers and land managers to earn carbon credits by storing carbon or reducing greenhouse gas emissions on the land. This new carbon price is expected to encourage the development of ecologically based, market-driven carbon plantings (Crossman, Bryan & Summers 2011). Because encroached communities are generally mixed-species stands, they are likely to offer additional biodiversity co-benefits that are not provided by monocultural plantings (Watson et al. 2000), and therefore may represent a superior C-storage strategy. It may be timely therefore to investigate alternatives to a predominantly grazing-centric view of encroachment, and to possibly embrace other land uses, some of which may be more profitable than pastoralism under new carbon economies.
Our research has shown that shrublands are associated with increases in measures of ecosystem function that are indicative of healthy dryland ecosystems, though sometimes neutral responses, and our results provide a strong ecological platform on which to reassess the value of shrublands. Ultimately, however, the prevailing land use is likely to depend on social systems and human decisions, and how society reconciles competing valuations of ecosystem services related to soil carbon, grazing and wildlife habitat. Developing useful models that couple natural systems with social systems is a key to informing this process.