Quantitative Assessment of Degradation Classes of Degraded Alpine Meadow (Heitutan)

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Chapter 2 Literature Review


A critical step in improving the way we manage grassland ecosystems is to take stock of their extent and their condition. These considerations underpin the long-term sustainable use of grassland. To date, no such comprehensive assessment of the state of grassland ecosystems on the QTP has been undertaken. In this chapter, the state of ecosystems and approaches to recovery/rehabilitation, the extent of rangeland/grassland degradation, primary controls upon degradation and the implications for their restoration and recovery on the QTP are reviewed and discussed. A comprehensive review of the relevant literature provides the evidence to demonstrate that rangeland degradation on the QTP is indeed a serious problem. The specific purposes of this chapter are: (i) to overview conceptual approaches that assess human impacts on ecosystems in the form of grassland degradation; (ii) to demonstrate the nature and extent of rangeland degradation on the QTP; (iii) to illustrate the variability of rangeland degradation severity with rangeland type and environmental settings, highlighting differing underlying causes of degradation; (iv) to prescribe targeted measures to effectively rehabilitate rangelands that have been degraded to varying degrees by different mechanisms; and (v) to review simulation modelling of the degradation and restoration of grassland so as to better understand its significance to management practice.

The world’s grassland and human impacts

Grassland ecosystems provide many ecological goods and services, including animal habitat, soil services, grazing resources for livestock and C storage (White et al. 2000; Lund, 2007). Suttie et al. (2005) reported that “grassland” may be defined as ground covered by vegetation dominated by grasses, with little or no tree cover. Grassland, rangeland and pasture are terms used in different parts of the world to refer to a similar set of environmental conditions wherein rangeland ecosystems produce biomass which can be grazed (Ren, 2000). Several studies have presented estimates of the extent of the world’s grassland areas, with estimates range from approximately 24 to 105 million km2, or 18 to 80 percent of the earth’s surface (White et al., 2000; Lund, 2007). However, estimates ofthe amount of the Earth’s land surface covered by rangelands vary considerably, in part reflecting both the range of definitions and data sources used to make them (Lund 2007). At present, information on range-/grassland biomes lags behind data for other ecosystems, so scientists and policy require additional data to address these issues in a comprehensive manner. Grassland/rangeland refers to expansive and mostly unimproved land that supports natural vegetation such as grasses or grass-like plants, including forbs and shrubs (Li et al., 2011).
Grasslands are found in every region of the world; Sub-Saharan Africa and Asia have the largest total area in grassland, 14.5 and 8.9 million km2 respectively (White et al., 2000). The five countries with the largest grassland area are Australia, the Russian Federation, China, the United States, and Canada. Dry land degradation may be the most severe anthropic modification of grassland ecosystems (Graetz, 1994). Domestic animals generate effects that may be either positive or negative, particularly in relation to different stocking densities and different grassland environments. Grasslands and populations of wild ungulates have coexisted for millions of years (White et al., 2000). Indeed, their emergence and subsequent adaptive radiation has been described as “among the most thoroughly documented evolutionary patterns in the fossil record” (Frank et al., 1998:519). A comparison of the grazing of grasslands by wild ungulates and domestic livestock underlines important transformations in grassland ecosystems. The primary goal of domestic animal husbandry is to maximize animal biomass, and a range of different tools (e.g., supplemental feeding, veterinary care) are used to achieve this, with the end result being a generally higher biomass of domestic as opposed to wild animals on grasslands (Frank et al., 1998). Oesterheld et al. (1992) found that, after controlling for primary productivity, the biomass of domestic animals is up to an order of magnitude higher than that of wild grazers. At high densities, grazing animals can change the floristic composition and structural characteristics of vegetation, reduce biodiversity, and increase soil erosion; in extreme situations, grazing may eliminate much vegetation cover (Evans, 1998). The extent to which these changes occur may depend not only on the number of livestock but also on the pattern of their grazing. As noted by Van de Koppel et al. (1997), many terrestrial grazing systems across the world have proved to be vulnerable to changes in grazing pressure. Increase in human population size has resulted in increasing grazing pressure by domestic animals in many parts of the world. Ultimately, this pressure will lead to environmental degradation (Cingolani et al., 2005). Grassland degradation can be defined as a process (including any change) in which rangeland quality declines to such a level that surface vegetation has been fragmented as a result of excessive human activities and/or unfavourable natural conditions (Li, 1997). Its manifestation includes initial lowering of rangeland productivity, fragmentation of grass cover, reduction in soil fertility, soil compaction, an increase in unpalatable grass species or a combination of all of these factors. Hence, rangeland degradation is detrimental to rangeland health and undermines animal husbandry. Rangeland degradation as directly affects the livelihood of pastoralists dependent on these ecosystems, but also indirectly affects people affected by its climatic, hydrologic and economic outcomes (Harris, 2010).
Grazing lands are possibly “the most degraded land use type in the world” (Papanastasis, 2009: 441). Destruction of vegetation and subsequent degradation have been related to increased livestock grazing pressure in grasslands in Africa, south-western USA, the Russian Federation and China (Van de Koppel et al., 1997). Grasslands in America and Australia cannot easily be restored simply by lowering the level of grazing animals. White et al. (2000) reported that indicators of soil condition showed that more than half of the grassland area in the world had some degree of soil degradation; nearly 49 percent were lightly to moderately degraded and over 5 percent of these grasslands were considered to be extremely degraded. Lund (2007) reported that the amount of the world’s grassland that is considered to be degraded ranges from 20% to 73%. The spatial patterns of herding of domestic animals by humans differ from the movements of wild ungulates and tend to result in more concentrated use of grasslands (Frank et al., 1998; McNaughton, 1993). Thus, the densities of domestic animals may be much higher than those of wild ungulates, and their grazing patterns may inhibit the recovery of defoliated grasses. Historical records suggest evidence of large-scale soil degradation in many areas of the world over the past 5,000 years (Scherr, 1999). White et al. (2000) suggest that the extent and severity of soil degradation has increased markedly in recent decades – such effects are especially pronounced on the QTP.

 Conceptual considerations in assessment of grassland degradation

‘Natural’ ecosystems provide humans with a wide-range of goods and services (White et al., 2000). At the same time, humanity has strongly influenced biogeochemical, hydrological,and ecological processes, from local to global scales (Folke et al., 2004). Land and grassland degradation are a world-wide problem (Kessler and Laban, 1994; Muller et al., 1998; Carrick and Krüger, 2007). The environmental damage resulting from population pressure, land use intensification and climate change, among many other factors, has been accompanied by progressive depletion of natural resources. The Global Assessment of Human-Induced Soil Degradation concluded that “approximately 23 percent of the world’s used terrestrial area was degraded: 38 percent was lightly degraded; 46 percent moderately degraded; and 16 percent strongly to extremely degraded” (Oldeman and van Lynden 1997: 430).
The idea that ecological systems (e.g. grassland) might take on alternative stable states or ‘basins of attraction’ (Beisner et al., 2003; Li et al., 2012b), with moves between these states occurring after specific abiotic and biotic thresholds are transgressed, has gained attention in restoration ecology (Suding et al. 2004; Suding and Hobbs, 2009). Positive feedbacks (e.g. between topography, soils and vegetation; sensu Suding and Hobbs, 2009) may hold ecosystems in a given stable state, with the potential for strong hysteresis effects (Groffman et al. 2006; Li et al., 2012b). Van de Koppel et al. (1997:355) noted that “a number of empirical studies indicate that plant-soil feedbacks are the dominant cause of catastrophic behaviour in many terrestrial grazing systems”. Although debate continues over their usefulness (Briske et al., 2003), there are increasingly sophisticated applications of these concepts to ecosystem management (Standish et al., 2009). Ecological thresholds arise where small changes in state variables result in large shifts in ecological conditions (Martin et al., 2009). In restoration science, differentiation of alternative stable states and associated threshold conditions can be used to frame management interventions, so long as they identify the underlying mechanisms that drive change (Bestelmeyer, 2006). For example, ecological
thresholds may reflect changes in vegetation and soils that result in shifts to degraded ecosystem states that are expensive or impossible to reverse. Identifying abiotic and biotic ecological thresholds, and their transgression, during ecosystem degradation is vital for the effective rehabilitation of degraded ecosystems (Hobbs and Cramer, 2008).Ideas stemming from the alternate stable state theory have often been applied to rangeland ecosystems (Bestelmeyer, 2006; Hobbs and Cramer, 2008). The effect of anthropogenic influence on rangeland degradation is conceptualized in Figure 2.1. The ecological trajectory of the original rangeland in the presence of human-related grazing is dictated by two forces, anthropogenic and climatic. Changes resulting from human activity may result in more immediate and profound changes to soil fertility and structure than that caused by climate change. The original and human affected grasslands in Figure 2.1 may be stable ecosystems (see Briske et al., 2003), but the moderately and severely degraded grasslands shown in Figure 2.1 may be unstable ecosystems because they might be extremely degraded (e.g. DeAngelis and Waterhouse, 1987). Except the moderate and severe degradations, the other four states of Figure 2.1 could be considered to represent alternative stable states (see Beisner et al., 2003).Rangeland degradation falls into reversible and ‘difficult to reverse’ categories (Li et al., 2011). Reversible degradation starts when the abiotic conditions of rangeland are still within its self-restoration capacity. At the initial stage human impact is considered minimal as both the ecological and the physical integrity of rangeland is retained. Any negative impacts on abiotic conditions are reversible in the sense that only the above ground biomass, structural composition of the plant community, and local soil condition (i.e. biotic factors) are affected. Rangelands at moderate and severe degradations retain sufficient resilience such that the recovery mechanisms are able to reverse degradational influences once external pressures are lessened or removed. Reversible degradation means the self-restoration capacity of the grassland is retained even when human disturbance is intense (Figure 2.1). By the time the system has moved to the extreme state, the capacity of the grassland ecosystem for self-restoration has been lost for an extended period (Zhou et al., 2008). After such a long time away from the ‘original’ state, the framework presented in Figure 2.1 suggests that the degradation process will be difficult to reverse. Rangeland abiotic conditions have been modified to such a level that there is little prospect of the self-recovery of dominant native species. Rangeland vegetation cannot quickly revert to its former state even if grazing intensity is reduced or when droughts end. Different measures are needed to rehabilitate rangelands that have been degraded by different mechanisms.
The framework outlined in Figure 2.1 can, potentially, be used in reverse to outline the prospects for restoration and management of degraded ecosystems based upon the degree of degradation, threshold relationships and recovery prospects (Figure 2.2). The underlying controls of grassland degradation are synthesized in the following section.

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List of Figures 
List of Tables 
Chapter 1 Introduction
1.1 Background
1.2 Objectives
1.3 Structure of the thesis
Chapter 2 Literature Review
2.1 Introduction
2.2 The world’s grassland and human impacts
2.3 Conceptual considerations in assessment of grassland degradation
2.4 State and degradation of grasslands on the QTP
2.5 Background considerations in the development of a simulation model to assess grassland degradation and restoration
2.6 Summary
Chapter 3 Regional Setting and ‘Heitutan’ Degradation
3.1 Introduction
3.2 Geographic features and ecological conditions of the Sanjiangyuan region
3.3 Cause of grassland degradation in the Sanjiangyuan
3.4 ‘Heitutan’ degradation in the Sanjiangyuan
3.5 Livestock grazing dynamics and small mammal outbreaks in the Sanjiangyuan region
3.6 Summary
Chapter 4 Quantitative Assessment of Degradation Classes of Degraded Alpine Meadow (Heitutan)
4.1 Introduction
4.2 Methods
4.3 Results and analysis
4.4 Discussion and degraded alpine meadow restoration
4.5 Summary
Chapter 5 Model Description
5.1 Introduction
5.2 Modelling approach and rationale/reason
5.3 Model structure and process
5.4 Model rules
5.5 Procedure modules
5.6 Summary
Chapter 6 Model Parameterisation and Sensitivity Analysis
6.1 Introduction
6.2 Model parameterisation
6.3 Sensitivity analysis
6.4 Model evaluation
6.5 Summary
Chapter 7 Model Results
7.1 Introduction
7.2 PFTs community dynamics and spatial patterns under different grazing regimes
7.3 Simulated dynamics of alpine meadow in the Sanjiangyuan
7.4 Time period required for Heitutan degradation .
7.5 Response of four PFTs and two types of ground to small mammal activity
7.6 Model experiments to restore Heitutan
7.7 Summary
Chapter 8 Discussion
8.1 Introduction
8.2 Community dynamics and spatial patterns of PFTs in alpine meadow
8.3 Simulated dynamics of alpine meadow on the QTP
8.4 Causes and timeframe of Heitutan formation 168
8.5 Restoration of Heitutan in alpine meadow
8.6 Threshold conditions during the formation and restoration of Heitutan
8.7 The resilience and rehabilitation of grazing-adapted alpine meadow ecosystems
8.8 Implications for sustainable economic development strategies in alpine meadow
8.9 Model limitations and effectiveness of the model
8.10 Summary
Chapter 9 Conclusions and Implications for Future Research
9.1 Conclusions
9.2 Implications for future research
The spatio-temporal dynamics of four plant-functional types (PFTs) in alpine meadow as affected by human disturbance, Sanjiangyuan region, China

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