Editing IPCC:AR6/SRCCL/Chapter-3 (section)

=== 3.6.1 SLM technologies and practices: On-the-ground actions ===

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A broad range of activities and measures can help avoid, reduce and reverse degradation across the dryland areas of the world. Many of these actions also contribute to climate change adaptation and mitigation, with further sustainable development co-benefits for poverty eradication and food security ( ''high confidence'' ) (Section 6.3). As preventing desertification is strongly preferable and more cost-effective than allowing land to degrade and then attempting to restore it (IPBES 2018b <sup>[[#fn:r973|973]]</sup> ; Webb et al. 2013 <sup>[[#fn:r974|974]]</sup> ), there is a growing emphasis on avoiding and reducing land degradation, following the Land Degradation Neutrality framework (Cowie et al. 2018 <sup>[[#fn:r975|975]]</sup> ; Orr et al. 2017 <sup>[[#fn:r976|976]]</sup> ) (Section 4.8.5).

An assessment is made of six activities and measures practicable across the biomes and anthromes of the dryland domain (Figure 3.10). This suite of actions is not exhaustive, but rather a set of activities that are particularly pertinent to global dryland ecosystems. They are not necessarily exclusive to drylands and are often implemented across a range of biomes and anthromes (Figure 3.10; for afforestation, see Section 3.7.2, Cross-Chapter Box 2 in Chapter 1, and Chapter 4 (Section 4.8.3)). The use of anthromes as a structuring element for response options is based on the essential role of interactions between social and ecological systems in driving desertification within coupled socio-ecological systems (Cherlet et al. 2018 <sup>[[#fn:r977|977]]</sup> ). The concept of the anthromes is defined in the Glossary and explored further in Chapters 1, 4 and 6.

The assessment of each action is twofold: firstly, to assess the ability of each action to address desertification and enhance climate change resilience, and secondly, to assess the potential impact of future climate change on the effectiveness of each action.

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'''Figure 3.10'''

<span id="the-typical-distribution-of-on-the-ground-actions-across-global-biomes-and-anthromes."></span>
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'''The typical distribution of on-the-ground actions across global biomes and anthromes.'''
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[[File:473088f6c6b462988728c7de805fdddd Figure-3.10-1024x511.jpg]]

The typical distribution of on-the-ground actions across global biomes and anthromes.

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==== 3.6.1.1 Integrated crop–soil–water management ====

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Forms of integrated cropland management have been practiced in drylands for thousands of years (Knörzer et al. 2009 <sup>[[#fn:r978|978]]</sup> ). Actions include planting a diversity of species including drought-resilient ecologically appropriate plants, reducing tillage, applying organic compost and fertiliser, adopting different forms of irrigation and maintaining vegetation and mulch cover. In the contemporary era, several of these actions have been adopted in response to climate change.

In terms of climate change ''adaptation'' , the resilience of agriculture to the impacts of climate change is strongly influenced by the underlying health and stability of soils as well as improvements in crop varieties, irrigation efficiency and supplemental irrigation, for example, through rainwater harvesting (medium evidence, high agreement) (Altieri et al. 2015 <sup>[[#fn:r979|979]]</sup> ; Amundson et al. 2015 <sup>[[#fn:r980|980]]</sup> ; Derpsch et al. 2010 <sup>[[#fn:r981|981]]</sup> ; Lal 1997 <sup>[[#fn:r982|982]]</sup> ; de Vries et al. 2012 <sup>[[#fn:r983|983]]</sup> ). Desertification often leads to a reduction in ground cover that in turn results in accelerated water and wind erosion and an associated loss of fertile topsoil that can greatly reduce the resilience of agriculture to climate change (medium evidence, high agreement) (Touré et al. 2019 <sup>[[#fn:r984|984]]</sup> ; Amundson et al. 2015 <sup>[[#fn:r985|985]]</sup> ; Borrelli et al. 2017 <sup>[[#fn:r986|986]]</sup> ; Pierre et al. 2017 <sup>[[#fn:r987|987]]</sup> ). Amadou et al. (2011) <sup>[[#fn:r988|988]]</sup> note that even a minimal cover of crop residues (100 kg ha– <sup>1</sup> ) can substantially decrease wind erosion.

Compared to conventional (flood or furrow) irrigation, drip irrigation methods are more efficient in supplying water to the plant root zone, resulting in lower water requirements and enhanced water use efficiency ( ''robust evidence, high agreement'' ) (Ibragimov et al. 2007 <sup>[[#fn:r989|989]]</sup> ; Narayanamoorthy 2010 <sup>[[#fn:r990|990]]</sup> ; Niaz et al. 2009 <sup>[[#fn:r991|991]]</sup> ). For example, in the rainfed area of Fetehjang, Pakistan, the adoption of drip methods reduced water usage by 67–68% during the production of tomato, cucumber and bell peppers, resulting in a 68–79% improvement in water use efficiency compared to previous furrow irrigation (Niaz et al. 2009 <sup>[[#fn:r992|992]]</sup> ). In India, drip irrigation reduced the amount of water consumed in the production of sugarcane by 44%, grapes by 37%, bananas by 29% and cotton by 45%, while enhancing yields by up to 29% (Narayanamoorthy 2010 <sup>[[#fn:r993|993]]</sup> ). Similarly, in Uzbekistan, drip irrigation increased the yield of cotton by 10–19% while reducing water requirements by 18–42% (Ibragimov et al. 2007 <sup>[[#fn:r994|994]]</sup> ).

A prominent response that addresses soil loss, health and cover is altering cropping methods. The adoption of intercropping (inter – and intra-row planting of companion crops) and relay cropping (temporally differentiated planting of companion crops) maintains soil cover over a larger fraction of the year, leading to an increase in production, soil nitrogen, species diversity and a decrease in pest abundance ( ''robust evidence, medium agreement'' ) (Altieri and Koohafkan 2008 <sup>[[#fn:r995|995]]</sup> ; Tanveer et al. 2017 <sup>[[#fn:r996|996]]</sup> ; Wilhelm and Wortmann 2004 <sup>[[#fn:r997|997]]</sup> ). For example, intercropping maize and sorghum with ''Desmodium'' (an insect repellent forage legume) and Brachiaria (an insect trapping grass), which is being promoted in drylands of East Africa, led to a two-to-three-fold increase in maize production and an 80% decrease in stem boring insects (Khan et al. 2014 <sup>[[#fn:r998|998]]</sup> ). In addition to changes in cropping methods, forms of agroforestry and shelterbelts are often used to reduce erosion and improve soil conditions (Section 3.7.2). For example, the use of tree belts of mixed species in northern China led to a reduction of surface wind speed and an associated reduction in soil temperature of up to 40% and an increase in soil moisture of up to 30% (Wang et al. 2008 <sup>[[#fn:r999|999]]</sup> ).

A further measure that can be of increasing importance under climate change is rainwater harvesting (RWH), including traditional ''zai'' (small basins used to capture surface runoff), earthen bunds and ridges (Nyamadzawo et al. 2013 <sup>[[#fn:r1001|1001]]</sup> ), ''fanya juus'' infiltration pits (Nyagumbo et al. 2019 <sup>[[#fn:r1002|1002]]</sup> ), contour stone bunds (Garrity et al. 2010 <sup>[[#fn:r1003|1003]]</sup> ) and semi-permeable stone bunds (often referred to by the French term ''digue filtrante'' ) (Taye et al. 2015 <sup>[[#fn:r1004|1004]]</sup> ). RWH increases the amount of water available for agriculture and livelihoods through the capture and storage of runoff, while at the same time reducing the intensity of peak flows following high-intensity rainfall events. It is therefore often highlighted as a practical response to dryness (i.e., long-term aridity and low seasonal precipitation) and rainfall variability, both of which are projected to become more acute over time in some dryland areas (Dile et al. 2013 <sup>[[#fn:r1005|1005]]</sup> ; Vohland and Barry 2009 <sup>[[#fn:r1006|1006]]</sup> ). For example, for drainage in Wadi Al-Lith, Saudi Arabia, the use of rainwater harvesting was suggested as a key climate change adaptation action (Almazroui et al. 2017 <sup>[[#fn:r1007|1007]]</sup> ). There is ''robust evidence'' and ''high agreement'' that the implementation of RWH systems leads to an increase in agricultural production in drylands (Biazin et al. 2012 <sup>[[#fn:r1008|1008]]</sup> ; Bouma and Wösten 2016 <sup>[[#fn:r1009|1009]]</sup> ; Dile et al. 2013 <sup>[[#fn:r1010|1010]]</sup> ). A global meta-analysis of changes in crop production due to the adoption of RWH techniques noted an average increase in yields of 78%, ranging from –28% to 468% (Bouma and Wösten 2016 <sup>[[#fn:r1011|1011]]</sup> ). Of particular relevance to climate change in drylands is that the relative impact of RWH on agricultural production generally increases with increasing dryness. Relative yield improvements due to the adoption of RWH were significantly higher in years with less than 330 mm rainfall, compared to years with more than 330 mm (Bouma and Wösten 2016 <sup>[[#fn:r1012|1012]]</sup> ). Despite delivering a clear set of benefits, there are some issues that need to be considered. The impact of RWH may vary at different temporal and spatial scales (Vohland and Barry 2009 <sup>[[#fn:r1013|1013]]</sup> ). At a plot scale, RWH structures may increase available water and enhance agricultural production, SOC and nutrient availability, yet at a catchment scale, they may reduce runoff to downstream uses (Meijer et al. 2013 <sup>[[#fn:r1014|1014]]</sup> ; Singh et al. 2012 <sup>[[#fn:r1015|1015]]</sup> ; Vohland and Barry 2009 <sup>[[#fn:r1016|1016]]</sup> ; Yosef and Asmamaw 2015 <sup>[[#fn:r1017|1017]]</sup> ). Inappropriate storage of water in warm climes can lead to an increase in water related diseases unless managed correctly, for example, schistosomiasis and malaria (Boelee et al. 2013 <sup>[[#fn:r1018|1018]]</sup> ).

Integrated crop–soil–water management may also deliver climate change ''mitigation'' benefits through avoiding, reducing and reversing the loss of SOC (Table 6.5). Approximately 20–30 Pg of SOC have been released into the atmosphere through desertification processes, for example, deforestation, overgrazing and conventional tillage (Lal 2004 <sup>[[#fn:r1019|1019]]</sup> ). Activities, such as those associated with conservation agriculture (minimising tillage, crop rotation, maintaining organic cover and planting a diversity of species), reduce erosion, improve water use efficiency and primary production, increase inflow of organic material and enhance SOC over time, contributing to climate change mitigation and adaptation ( ''high confidence'' ) (Plaza-Bonilla et al. 2015 <sup>[[#fn:r1020|1020]]</sup> ; Lal 2015 <sup>[[#fn:r1021|1021]]</sup> ; Srinivasa Rao et al. 2015 <sup>[[#fn:r1022|1022]]</sup> ; Sombrero and de Benito 2010 <sup>[[#fn:r1023|1023]]</sup> ). Conservation agriculture practices also lead to increases in SOC ( ''medium confidence'' ). However, sustained carbon sequestration is dependent on net primary productivity and on the availability of crop-residues that may be relatively limited and often consumed by livestock or used elsewhere in dryland contexts (Cheesman et al. 2016 <sup>[[#fn:r1024|1024]]</sup> ; Plaza-Bonilla et al. 2015 <sup>[[#fn:r1025|1025]]</sup> ). For this reason, expected rates of carbon sequestration following changes in agricultural practices in drylands are relatively low (0.04–0.4 tC ha <sup>–1</sup> ) and it may take a protracted period of time, even several decades, for carbon stocks to recover if lost ( ''medium confidence'' ) (Farage et al. 2007 <sup>[[#fn:r1026|1026]]</sup> ; Hoyle et al. 2013 <sup>[[#fn:r1027|1027]]</sup> ; Lal 2004 <sup>[[#fn:r1028|1028]]</sup> ). This long recovery period enforces the rationale for prioritising the avoidance and reduction of land degradation and loss of C, in addition to restoration activities.

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==== 3.6.1.2 Grazing and fire management in drylands ====

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Rangeland management systems such as sustainable grazing approaches and re-vegetation increase rangeland productivity ( ''high confidence'' ) (Table 6.5). Open grassland, savannah and woodland are home to the majority of world’s livestock production (Safriel et al. 2005 <sup>[[#fn:r1029|1029]]</sup> ). Within these drylands areas, prevailing grazing and fire regimes play an important role in shaping the relative abundance of trees versus grasses (Scholes and Archer 1997 <sup>[[#fn:r1030|1030]]</sup> ; Staver et al. 2011 <sup>[[#fn:r1031|1031]]</sup> ; Stevens et al. 2017 <sup>[[#fn:r1032|1032]]</sup> ), as well as the health of the grass layer in terms of primary production, species richness and basal cover (the propotion of the plant that is in the soil) (Plaza-Bonilla et al. 2015 <sup>[[#fn:r1033|1033]]</sup> ; Short et al. 2003 <sup>[[#fn:r1034|1034]]</sup> ). This in turn influences levels of soil erosion, soil nutrients, secondary production and additional ecosystem services (Divinsky et al. 2017 <sup>[[#fn:r1035|1035]]</sup> ; Pellegrini et al. 2017 <sup>[[#fn:r1036|1036]]</sup> ). A further set of drivers, including soil type, annual rainfall and changes in atmospheric CO <sub>2</sub> may also define observed rangeland structure and composition (Devine et al. 2017 <sup>[[#fn:r1037|1037]]</sup> ; Donohue et al. 2013 <sup>[[#fn:r1038|1038]]</sup> ), but the two principal factors that pastoralists can manage are grazing and fire, by altering their frequency, type and intensity.

The impact of grazing and fire regimes on biodiversity, soil nutrients, primary production and further ecosystem services is not constant and varies between locations (Divinsky et al. 2017 <sup>[[#fn:r1039|1039]]</sup> ; Fleischner 1994 <sup>[[#fn:r1040|1040]]</sup> ; van Oijen et al. 2018 <sup>[[#fn:r1041|1041]]</sup> ). Trade-offs may therefore need to be considered to ensure that rangeland diversity and production are resilient to climate change (Plaza-Bonilla et al. 2015 <sup>[[#fn:r1042|1042]]</sup> ; van Oijen et al. 2018 <sup>[[#fn:r1043|1043]]</sup> ). In certain locations, even light to moderate grazing has led to a significant decrease in the occurrence of particular species, especially forbs (O’Connor et al. 2011 <sup>[[#fn:r1044|1044]]</sup> ; Scott-shaw and Morris 2015 <sup>[[#fn:r1045|1045]]</sup> ). In other locations, species richness is only significantly impacted by heavy grazing and is able to withstand light to moderate grazing (Divinsky et al. 2017 <sup>[[#fn:r1046|1046]]</sup> ). A context specific evaluation of how grazing and fire impact particular species may therefore be required to ensure the persistence of target species over time (Marty 2005 <sup>[[#fn:r1047|1047]]</sup> ). A similar trade-off may need to be considered between soil carbon sequestration and livestock production. As noted by Plaza-Bonilla et al. (2015) <sup>[[#fn:r1048|1048]]</sup> increasing grazing pressure has been found to increase SOC stocks in some locations, and decrease them in others. Where it has led to a decrease in soil carbon stocks, for example in Mongolia (Han et al. 2008 <sup>[[#fn:r1049|1049]]</sup> ) and Ethiopia (Bikila et al. 2016 <sup>[[#fn:r1050|1050]]</sup> ), trade-offs between carbon sequestration and the value of livestock to local livelihoods need be considered.

Although certain herbaceous species may be unable to tolerate grazing pressure, a complete lack of grazing or fire may not be desired in terms of ecosystems health. It can lead to a decrease in basal cover and the accumulation of moribund, unpalatable biomass that inhibits primary production (Manson et al. 2007 <sup>[[#fn:r1051|1051]]</sup> ; Scholes 2009 <sup>[[#fn:r1052|1052]]</sup> ). The utilisation of the grass sward through light to moderate grazing stimulates the growth of biomass and basal cover, and allows water services to be sustained over time (Papanastasis et al. 2017 <sup>[[#fn:r1053|1053]]</sup> ; Scholes 2009 <sup>[[#fn:r1054|1054]]</sup> ). Even moderate to heavy grazing in periods of higher rainfall may be sustainable, but constant heavy grazing during dry periods, and especially droughts, can lead to a reduction in basal cover, SOC, biological soil crusts, ecosystem services and an accelerated erosion ( ''high agreement, robust evidence'' ) (Archer et al. 2017 <sup>[[#fn:r1055|1055]]</sup> ; Conant and Paustian 2003 <sup>[[#fn:r1056|1056]]</sup> ; D’Odorico et al. 2013 <sup>[[#fn:r1057|1057]]</sup> ; Geist and Lambin 2004 <sup>[[#fn:r1058|1058]]</sup> ; Havstad et al. 2006 <sup>[[#fn:r1059|1059]]</sup> ; Huang et al. 2007 <sup>[[#fn:r1060|1060]]</sup> ; Manzano and Návar 2000 <sup>[[#fn:r1061|1061]]</sup> ; Pointing and Belnap 2012 <sup>[[#fn:r1062|1062]]</sup> ; Weber et al. 2016 <sup>[[#fn:r1063|1063]]</sup> ). For this reason, the inclusion of drought forecasts and contingency planning in grazing and fire management programmes is crucial to avoid desertification (Smith and Foran 1992 <sup>[[#fn:r1064|1064]]</sup> ; Torell et al. 2010 <sup>[[#fn:r1065|1065]]</sup> ). It is an important component of avoiding and reducing early degradation. Although grasslands systems may be relatively resilient and can often recover from a moderately degraded state (Khishigbayar et al. 2015 <sup>[[#fn:r1066|1066]]</sup> ; Porensky et al. 2016 <sup>[[#fn:r1067|1067]]</sup> ), if a tipping point has been exceeded, restoration to a historic state may not be economical or ecologically feasible (D’Odorico et al. 2013 <sup>[[#fn:r1068|1068]]</sup> ).

Together with livestock management (Table 6.5), the use of fire is an integral part of rangeland management, which can be applied to remove moribund and unpalatable forage, exotic weeds and woody species (Archer et al. 2017 <sup>[[#fn:r1069|1069]]</sup> ). Fire has less of an effect on SOC and soil nutrients in comparison to grazing (Abril et al. 2005 <sup>[[#fn:r1070|1070]]</sup> ), yet elevated fire frequency has been observed to lead to a decrease in soil carbon and nitrogen (Abril et al. 2005 <sup>[[#fn:r1071|1071]]</sup> ; Bikila et al. 2016 <sup>[[#fn:r1072|1072]]</sup> ; Bird et al. 2000 <sup>[[#fn:r1073|1073]]</sup> ; Pellegrini et al. 2017 <sup>[[#fn:r1074|1074]]</sup> ). Although the impact of climate change on fire frequency and intensity may not be clear due to its differing impact on fuel accumulation, suitable weather conditions and sources of ignition (Abatzoglou et al. 2018 <sup>[[#fn:r1075|1075]]</sup> ; Littell et al. 2018 <sup>[[#fn:r1076|1076]]</sup> ; Moritz et al. 2012 <sup>[[#fn:r1077|1077]]</sup> ), there is an increasing use of prescribed fire to address several global change phenomena, for example, the spread of invasive species and bush encroachment, as well as the threat of intense runaway fires (Fernandes et al. 2013 <sup>[[#fn:r1078|1078]]</sup> ; McCaw 2013 <sup>[[#fn:r1079|1079]]</sup> ; van Wilgen et al. 2010 <sup>[[#fn:r1080|1080]]</sup> ). Cross-Chapter Box 3 in Chapter 2 provides a further review of the interaction between fire and climate change.

There is often much emphasis on reducing and reversing the degradation of rangelands due to the wealth of benefits they provide, especially in the context of assisting dryland communities to adapt to climate change (Webb et al. 2017 <sup>[[#fn:r1081|1081]]</sup> ; Woollen et al. 2016 <sup>[[#fn:r1082|1082]]</sup> ). The emerging concept of ecosystem-based adaptation has highlighted the broad range of important ecosystem services that healthy rangelands can provide in a resilient manner to local residents and downstream economies (Kloos and Renaud 2016 <sup>[[#fn:r1083|1083]]</sup> ; Reid et al. 2018 <sup>[[#fn:r1084|1084]]</sup> ). In terms of climate change mitigation, the contribution of rangelands, woodland and sub-humid dry forest (e.g., Miombo woodland in south-central Africa) is often undervalued due to relatively low carbon stocks per hectare. Yet due to their sheer extent, the amount of carbon sequestered in these ecosystems is substantial and can make a valuable contribution to climate change mitigation (Lal 2004 <sup>[[#fn:r1085|1085]]</sup> ; Pelletier et al. 2018 <sup>[[#fn:r1086|1086]]</sup> ).

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==== 3.6.1.3 Clearance of bush encroachment ====

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The encroachment of open grassland and savannah ecosystems by woody species has occurred for at least the past 100 years (Archer et al. 2017 <sup>[[#fn:r1087|1087]]</sup> ; O’Connor et al. 2014 <sup>[[#fn:r1088|1088]]</sup> ; Schooley et al. 2018 <sup>[[#fn:r1089|1089]]</sup> ). Dependent on the type and intensity of encroachment, it may lead to a net loss of ecosystem services and be viewed as a form of desertification (Dougill et al. 2016 <sup>[[#fn:r1090|1090]]</sup> ; O’Connor et al. 2014 <sup>[[#fn:r1091|1091]]</sup> ). However, there are circumstances where bush encroachment may lead to a net increase in ecosystem services, especially at intermediate levels of encroachment, where the ability of the landscape to produce fodder for livestock is retained, while the production of wood and associated products increases (Eldridge et al. 2011 <sup>[[#fn:r1092|1092]]</sup> ; Eldridge and Soliveres 2014 <sup>[[#fn:r1093|1093]]</sup> ). This may be particularly important in regions such as southern Africa and India where over 65% of rural households depend on fuelwood from surrounding landscapes as well as livestock production (Komala and Prasad 2016 <sup>[[#fn:r1094|1094]]</sup> ; Makonese et al. 2017 <sup>[[#fn:r1095|1095]]</sup> ; Shackleton and Shackleton 2004 <sup>[[#fn:r1096|1096]]</sup> ).

This variable relationship between the level of encroachment, carbon stocks, biodiversity, provision of water and pastoral value (Eldridge and Soliveres 2014 <sup>[[#fn:r1097|1097]]</sup> ) can present a conundrum to policymakers, especially when considering the goals of three Rio Conventions: UNFCCC, UNCCD and UNCBD. Clearing intense bush encroachment may improve species diversity, rangeland productivity, the provision of water and decrease desertification, thereby contributing to the goals of the UNCBD and UNCCD as well as the adaptation aims of the UNFCCC. However, it would lead to the release of biomass carbon stocks into the atmosphere and potentially conflict with the mitigation aims of the UNFCCC.

For example, Smit et al. (2015) <sup>[[#fn:r1098|1098]]</sup> observed an average increase in above-ground woody carbon stocks of 44 tC ha <sup>–1</sup> in savannahs in northern Namibia. However, since bush encroachment significantly inhibited livestock production, there are often substantial efforts to clear woody species (Stafford-Smith et al. 2017 <sup>[[#fn:r1099|1099]]</sup> ). Namibia has a national programme, currently in its early stages, aimed at clearing woody species through mechanical measures (harvesting of trees) as well as the application of arboricides (Smit et al. 2015 <sup>[[#fn:r1100|1100]]</sup> ). However, the long-term success of clearance and subsequent improved fire and grazing management remains to be evaluated, especially restoration back towards an ‘original open grassland state’. For example, in northern Namibia, the rapid reestablishment of woody seedlings has raised questions about whether full clearance and restoration is possible (Smit et al. 2015 <sup>[[#fn:r1101|1101]]</sup> ). In arid landscapes, the potential impact of elevated atmospheric CO <sub>2</sub> (Donohue et al. 2013 <sup>[[#fn:r1102|1102]]</sup> ; Kgope et al. 2010 <sup>[[#fn:r1103|1103]]</sup> ) and opportunity to implement high-intensity fires that remove woody species and maintain rangelands in an open state has been questioned (Bond and Midgley 2000 <sup>[[#fn:r1104|1104]]</sup> ). If these drivers of woody plant encroachment cannot be addressed, a new form of ‘emerging ecosystem’ (Milton 2003 <sup>[[#fn:r1105|1105]]</sup> ) may need to be explored that includes both improved livestock and fire management as well as the utilisation of biomass as a long-term commodity and source of revenue (Smit et al. 2015 <sup>[[#fn:r1106|1106]]</sup> ). Initial studies in Namibia and South Africa (Stafford-Smith et al. 2017 <sup>[[#fn:r1107|1107]]</sup> ) indicate that there may be good opportunity to produce sawn timber, fencing poles, fuelwood and commercial energy, but factors such as the cost of transport can substantially influence the financial feasibility of implementation.

The benefit of proactive management that prevents land from being degraded (altering grazing systems or treating bush encroachment at early stages before degradation has been initiated) is more cost-effective in the long term and adds more resistance to climate change than treating lands after degradation has occurred (Webb et al. 2013 <sup>[[#fn:r1108|1108]]</sup> ; Weltz and Spaeth 2012 <sup>[[#fn:r1109|1109]]</sup> ). The challenge is getting producers to alter their management paradigm from short-term objectives to long-term objectives.

<div id="section-3-6-1-4-combating-sand-and-dust-storms-through-sand-dune-stabilisation"></div>

<span id="combating-sand-and-dust-storms-through-sand-dune-stabilisation"></span>
==== 3.6.1.4 Combating sand and dust storms through sand dune stabilisation ====

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Dust and sand storms have a considerable impact on natural and human systems (Sections 3.4.1 and 3.4.2). Application of sand dune stabilisation techniques contributes to reducing sand and dust storms ( ''high confidence'' ). Using a number of methods, sand dune stabilisation aims to avoid and reduce the occurrence of dust and sand storms (Mainguet and Dumay 2011 <sup>[[#fn:r1110|1110]]</sup> ). Mechanical techniques include building palisades to prevent the movement of sand and reduce sand deposits on infrastructure. Chemical methods include the use of calcium bentonite or using silica gel to fix mobile sand (Aboushook et al. 2012 <sup>[[#fn:r1111|1111]]</sup> ; Rammal and Jubair 2015 <sup>[[#fn:r1112|1112]]</sup> ). Biological methods include the use of mulch to stabilise surfaces (Sebaa et al. 2015 <sup>[[#fn:r1113|1113]]</sup> ; Yu et al. 2004 <sup>[[#fn:r1114|1114]]</sup> ) and establishing permanent plant cover using pasture species that improve grazing at the same time (Abdelkebir and Ferchichi 2015 <sup>[[#fn:r1115|1115]]</sup> ; Zhang et al. 2015 <sup>[[#fn:r1116|1116]]</sup> ) (Section 3.7.1.3). When the dune is stabilised, woody perennials are introduced that are selected according to climatic and ecological conditions (FAO 2011 <sup>[[#fn:r1117|1117]]</sup> ). For example, such re-vegetation processes have been implemented on the shifting dunes of the Tengger Desert in northern China leading to the stabilisation of sand and the sequestration of up to 10 tC ha <sup>–1</sup> over a period of 55 years (Yang et al. 2014 <sup>[[#fn:r1118|1118]]</sup> ).

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==== 3.6.1.5 Use of halophytes for the re-vegetation of saline lands ====

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Soil salinity and sodicity can severely limit the growth and productivity of crops (Jan et al. 2017 <sup>[[#fn:r1119|1119]]</sup> ) and lead to a decrease in available arable land. Leaching and drainage provides a possible solution, but can be prohibitively expensive. An alternative, more economical option, is the growth of halophytes (plants that are adapted to grow under highly saline conditions) that allow saline land to be used in a productive manner (Qadir et al. 2000 <sup>[[#fn:r1120|1120]]</sup> ). The biomass produced can be used as forage, food, feed, essential oils, biofuel, timber, or fuelwood (Chughtai et al. 2015 <sup>[[#fn:r1121|1121]]</sup> ; Mahmood et al. 2016 <sup>[[#fn:r1122|1122]]</sup> ; Sharma et al. 2016 <sup>[[#fn:r1123|1123]]</sup> ). A further co-benefit is the opportunity to mitigate climate change through the enhancement of terrestrial carbon stocks as land is re-vegetated (Dagar et al. 2014 <sup>[[#fn:r1124|1124]]</sup> ; Wicke et al. 2013 <sup>[[#fn:r1125|1125]]</sup> ). The combined use of salt-tolerant crops, improved irrigation practices, chemical remediation measures and appropriate mulch and compost is effective in reducing the impact of secondary salinisation ( ''medium confidence'' ).

In Pakistan, where about 6.2 Mha of agricultural land is affected by salinity, pioneering work on utilising salt-tolerant plants for the re-vegetation of saline lands (biosaline agriculture) was done in the early 1970s (NIAB 1997 <sup>[[#fn:r1796|1796]]</sup> ). A number of local and exotic varieties were initially screened for salt tolerance in lab – and greenhouse-based studies, and then distributed to similar saline areas (Ashraf et al. 2010 <sup>[[#fn:r1126|1126]]</sup> ). These included tree species ( ''Acacia ampliceps, Acacia nilotica, Eucalyptus camaldulensis, Prosopis juliflora, Azadirachta indica'' ) (Awan and Mahmood 2017 <sup>[[#fn:r1127|1127]]</sup> ), forage plants ( ''Leptochloa fusca, Sporobolus'' ''arabicus, Brachiaria mutica, Echinochloa'' sp., ''Sesbania'' and ''Atriplex'' spp.) and crop species including varieties of barley ( ''Hordeum vulgare'' ), cotton, wheat ( ''Triticum aestivum'' ) and ''Brassica'' spp. (Mahmood et al. 2016 <sup>[[#fn:r1128|1128]]</sup> ) as well as fruit crops in the form of date palm ( ''Phoenix dactylifera'' ) that has high salt tolerance with no visible adverse effects on seedlings (Yaish and Kumar 2015 <sup>[[#fn:r1129|1129]]</sup> ; Al-Mulla et al. 2013 <sup>[[#fn:r1130|1130]]</sup> ; Alrasbi et al. 2010 <sup>[[#fn:r1131|1131]]</sup> ). Pomegranate ( ''Punica granatum L.'' ) is another fruit crop of moderate to high salt tolerance. Through regulating growth form and nutrient balancing, it can maintain water content, chlorophyll fluorescence and enzyme activity at normal levels (Ibrahim 2016 <sup>[[#fn:r1132|1132]]</sup> ; Okhovatian-Ardakani et al. 2010 <sup>[[#fn:r1133|1133]]</sup> ).

In India and elsewhere, tree species including ''Prosopis juliflora, Dalbergia sissoo'' , and ''Eucalyptus tereticornis'' have been used to re-vegetate saline land. Certain biofuel crops in the form of ''Ricinus communis'' (Abideen et al. 2014 <sup>[[#fn:r1134|1134]]</sup> ), ''Euphorbia antisyphilitica'' (Dagar et al. 2014 <sup>[[#fn:r1135|1135]]</sup> ), ''Karelinia caspia'' (Akinshina et al. 2016 <sup>[[#fn:r1797|1797]]</sup> ) and ''Salicornia'' spp. (Sanandiya and Siddhanta 2014 <sup>[[#fn:r1136|1136]]</sup> ) are grown in saline areas, and ''Panicum turgidum'' (Koyro et al. 2013 <sup>[[#fn:r1137|1137]]</sup> ) and ''Leptochloa fusca'' (Akhter et al. 2003 <sup>[[#fn:r1138|1138]]</sup> ) have been grown as fodder crop on degraded soils with brackish water. In China, intense efforts are being made on the use of halophytes (Sakai et al. 2012 <sup>[[#fn:r1139|1139]]</sup> ; Wang et al. 2018 <sup>[[#fn:r1140|1140]]</sup> ). These examples reveal that there is great scope for saline areas to be used in a productive manner through the utilisation of halophytes. The most productive species often have yields equivalent to conventional crops, at salinity levels matching even that of seawater.

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