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

== 3.6 Responses to desertification under climate change ==

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Achieving sustainable development of dryland livelihoods requires avoiding dryland degradation through SLM and restoring and rehabilitating the degraded drylands due to their potential wealth of ecosystem benefits and importance to human livelihoods and economies (Thomas 2008 <sup>[[#fn:r965|965]]</sup> ). A broad suite of on-the-ground response measures exists to address desertification (Scholes 2009 <sup>[[#fn:r966|966]]</sup> ), be it in the form of improved fire and grazing management, the control of erosion; integrated crop, soil and water management, among others (Liniger and Critchley 2007 <sup>[[#fn:r967|967]]</sup> ; Scholes 2009 <sup>[[#fn:r968|968]]</sup> ). These actions are part of the broader context of dryland development and long-term SLM within coupled socio-economic systems (Reynolds et al. 2007 <sup>[[#fn:r969|969]]</sup> ; Stringer et al. 2017 <sup>[[#fn:r970|970]]</sup> ; Webb et al. 2017 <sup>[[#fn:r971|971]]</sup> ). Many of these response options correspond to those grouped under ‘land transitions’ in the IPCC Special Report on Global Warming of 1.5°C (Coninck et al. 2018 <sup>[[#fn:r972|972]]</sup> ) (Table 6.4). It is therefore recognised that such actions require financial, institutional and policy support for their wide-scale adoption and sustainability over time (Sections 3.6.3, 4.8.5 and 6.4.4).

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=== 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> ).

<div id="section-3-6-1-3-clearance-of-bush-encroachment"></div>

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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 ====

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

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> ).

<div id="section-3-6-1-5-use-of-halophytes-for-the-re-vegetation-of-saline-lands"></div>

<span id="use-of-halophytes-for-the-re-vegetation-of-saline-lands"></span>
==== 3.6.1.5 Use of halophytes for the re-vegetation of saline lands ====

<div id="section-3-6-1-5-use-of-halophytes-for-the-re-vegetation-of-saline-lands-block-1"></div>

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.

<span id="socio-economic-responses"></span>
=== 3.6.2 Socio-economic responses ===

<div id="section-3-6-2-socio-economic-responses-block-1"></div>

Socio-economic and policy responses are often crucial in enhancing the adoption of SLM practices (Cordingley et al. 2015 <sup>[[#fn:r1143|1143]]</sup> ; Fleskens and Stringer 2014 <sup>[[#fn:r1144|1144]]</sup> ; Nyanga et al. 2016 <sup>[[#fn:r1145|1145]]</sup> ) and for assisting agricultural households to diversify their sources of income (Barrett et al. 2017 <sup>[[#fn:r1146|1146]]</sup> ; Shiferaw and Djido 2016 <sup>[[#fn:r1147|1147]]</sup> ). Technology and socio-economic responses are not independent, but continuously interact.

<div id="section-3-6-2-1-socio-economic-responses-for-combating-desertification-under-climate-change"></div>

<span id="socio-economic-responses-for-combating-desertification-under-climate-change"></span>
==== 3.6.2.1 Socio-economic responses for combating desertification under climate change ====

<div id="section-3-6-2-1-socio-economic-responses-for-combating-desertification-under-climate-change-block-1"></div>

Desertification limits the choice of potential climate change mitigation and adaptation response options by reducing climate change adaptive capacities. Furthermore, many additional factors, for example, a lack of access to markets or insecurity of land tenure, hinder the adoption of SLM. These factors are largely beyond the control of individuals or local communities and require broader policy interventions (Section 3.6.3). Nevertheless, local collective action and ILK are still crucial to the ability of households to respond to the combined challenge of climate change and desertification. Raising awareness, capacity building and development to promote collective action and indigenous and local knowledge contribute to avoiding, reducing and reversing desertification under changing climate.

'''The use of indigenous and local knowledge''' enhances the success of SLM and its ability to address desertification (Altieri and Nicholls 2017 <sup>[[#fn:r1148|1148]]</sup> ; Engdawork and Bork 2016 <sup>[[#fn:r1149|1149]]</sup> ). Using indigenous and local knowledge for combating desertification could contribute to climate change adaptation strategies (Belfer et al. 2017 <sup>[[#fn:r1150|1150]]</sup> ; Codjoe et al. 2014 <sup>[[#fn:r1151|1151]]</sup> ; Etchart 2017 <sup>[[#fn:r1152|1152]]</sup> ; Speranza et al. 2010 <sup>[[#fn:r1153|1153]]</sup> ; Makondo and Thomas 2018 <sup>[[#fn:r1154|1154]]</sup> ; Maldonado et al. 2016 <sup>[[#fn:r1155|1155]]</sup> ; Nyong et al. 2007 <sup>[[#fn:r1156|1156]]</sup> ). There are abundant examples of how indigenous and local knowledge, which are an important part of broader agroecological knowledge (Altieri 2018 <sup>[[#fn:r1157|1157]]</sup> ), have allowed livelihood systems in drylands to be maintained despite environmental constraints. An example is the numerous traditional water harvesting techniques that are used across the drylands to adapt to dry spells and climate change. These include creating planting pits ( ''zai, ngoro'' ) and micro-basins, contouring hill slopes and terracing (Biazin et al. 2012 <sup>[[#fn:r1158|1158]]</sup> ) (Section 3.6.1). Traditional ''ndiva'' water harvesting systems in Tanzania enable the capture of runoff water from highland areas to downstream community-managed micro-dams for subsequent farm delivery through small-scale canal networks (Enfors and Gordon 2008 <sup>[[#fn:r1159|1159]]</sup> ). A further example are pastoralist communities located in drylands who have developed numerous methods to sustainably manage rangelands. Pastoralist communities in Morocco developed the ''agdal'' system of seasonally alternating use of rangelands to limit overgrazing (Dominguez 2014 <sup>[[#fn:r1160|1160]]</sup> ) as well as to manage forests in the Moroccan High Atlas Mountains (Auclair et al. 2011 <sup>[[#fn:r1161|1161]]</sup> ). Across the Arabian Peninsula and North Africa, a rotational grazing system, ''hema'' , was historically practiced by the Bedouin communities (Hussein 2011 <sup>[[#fn:r1162|1162]]</sup> ; Louhaichi and Tastad 2010 <sup>[[#fn:r1163|1163]]</sup> ). The Beni-Amer herders in the Horn of Africa have developed complex livestock breeding and selection systems (Fre 2018 <sup>[[#fn:r1164|1164]]</sup> ). Although well adapted to resource-sparse dryland environments, traditional practices are currently not able to cope with increased demand for food and environmental changes (Enfors and Gordon 2008 <sup>[[#fn:r1165|1165]]</sup> ; Engdawork and Bork 2016 <sup>[[#fn:r1166|1166]]</sup> ). Moreover, there is ''robust evidence'' documenting the marginalisation or loss of indigenous and local knowledge (Dominguez 2014 <sup>[[#fn:r1167|1167]]</sup> ; Fernández-Giménez and Fillat Estaque 2012 <sup>[[#fn:r1168|1168]]</sup> ; Hussein 2011 <sup>[[#fn:r1169|1169]]</sup> ; Kodirekkala 2017 <sup>[[#fn:r1170|1170]]</sup> ; Moreno-Calles et al. 2012 <sup>[[#fn:r1171|1171]]</sup> ). Combined use of indigenous and local knowledge and new SLM technologies can contribute to raising resilience to the challenges of climate change and desertification (high confidence) (Engdawork and Bork 2016 <sup>[[#fn:r1172|1172]]</sup> ; Guzman et al. 2018 <sup>[[#fn:r1173|1173]]</sup> ).

'''Collective action''' has the potential to contribute to SLM and climate change adaptation ( ''medium confidence'' ) (Adger 2003 <sup>[[#fn:r1174|1174]]</sup> ; Engdawork and Bork 2016 <sup>[[#fn:r1175|1175]]</sup> ; Eriksen and Lind 2009 <sup>[[#fn:r1176|1176]]</sup> ; Ostrom 2009 <sup>[[#fn:r1177|1177]]</sup> ; Rodima-Taylor et al. 2012 <sup>[[#fn:r1178|1178]]</sup> ). Collective action is a result of social capital. Social capital is divided into structural and cognitive forms: structural corresponding to strong networks (including outside one’s immediate community); and cognitive encompassing mutual trust and cooperation within communities (van Rijn et al. 2012 <sup>[[#fn:r1179|1179]]</sup> ; Woolcock and Narayan 2000 <sup>[[#fn:r1180|1180]]</sup> ). Social capital is more important for economic growth in settings with weak formal institutions, and less so in those with strong enforcement of formal institutions (Ahlerup et al. 2009 <sup>[[#fn:r1181|1181]]</sup> ). There are cases throughout the drylands showing that community by-laws and collective action successfully limited land degradation and facilitated SLM (Ajayi et al. 2016 <sup>[[#fn:r1182|1182]]</sup> ; Infante 2017 <sup>[[#fn:r1183|1183]]</sup> ; Kassie et al. 2013 <sup>[[#fn:r1184|1184]]</sup> ; Nyangena 2008 <sup>[[#fn:r1185|1185]]</sup> ; Willy and Holm-Müller 2013 <sup>[[#fn:r1186|1186]]</sup> ; Wossen et al. 2015 <sup>[[#fn:r1187|1187]]</sup> ). However, there are also cases when they did not improve SLM where they were not strictly enforced (Teshome et al. 2016 <sup>[[#fn:r1188|1188]]</sup> ). Collective action for implementing responses to dryland degradation is often hindered by local asymmetric power relations and ‘elite capture’ (Kihiu 2016 <sup>[[#fn:r1189|1189]]</sup> ; Stringer et al. 2007 <sup>[[#fn:r1190|1190]]</sup> ). This illustrates that different levels and types of social capital result in different levels of collective action. In a sample of East, West and southern African countries, structural social capital in the form of access to networks outside one’s own community was suggested to stimulate the adoption of agricultural innovations, whereas cognitive social capital, associated with inward-looking community norms of trust and cooperation, was found to have a negative relationship with the adoption of agricultural innovations (van Rijn et al. 2012 <sup>[[#fn:r1191|1191]]</sup> ). The latter is indirectly corroborated by observations of the impact of community-based rangeland management organisations in Mongolia. Although levels of cognitive social capital did not differ between them, communities with strong links to outside networks were able to apply more innovative rangeland management practices in comparison to communities without such links (Ulambayar et al. 2017 <sup>[[#fn:r1192|1192]]</sup> ).

'''Farmer-led innovations.''' Agricultural households are not just passive adopters of externally developed technologies, but are active experimenters and innovators (Reij and Waters-Bayer 2001 <sup>[[#fn:r1193|1193]]</sup> ; Tambo and Wünscher 2015 <sup>[[#fn:r1194|1194]]</sup> ; Waters-Bayer et al. 2009 <sup>[[#fn:r1195|1195]]</sup> ). SLM technologies co-generated through direct participation of agricultural households have higher chances of being accepted by them ( ''medium confidence'' ) (Bonney et al. 2016 <sup>[[#fn:r1196|1196]]</sup> ; Vente et al. 2016 <sup>[[#fn:r1197|1197]]</sup> ). Usually farmer-driven innovations are more frugal and better adapted to their resource scarcities than externally introduced technologies (Gupta et al. 2016 <sup>[[#fn:r1198|1198]]</sup> ). Farmer-to-farmer sharing of their own innovations and mutual learning positively contribute to higher technology adoption rates (Dey et al. 2017 <sup>[[#fn:r1199|1199]]</sup> ). This innovative ability can be given a new dynamism by combining it with emerging external technologies. For example, emerging low-cost phone applications (‘apps’) that are linked to soil and water monitoring sensors can provide farmers with previously inaccessible information and guidance (Cornell et al. 2013 <sup>[[#fn:r1200|1200]]</sup> ; Herrick et al. 2017 <sup>[[#fn:r1201|1201]]</sup> ; McKinley et al. 2017 <sup>[[#fn:r1202|1202]]</sup> ; Steger et al. 2017 <sup>[[#fn:r1203|1203]]</sup> ).

Currently, the adoption of SLM practices remains insufficient to address desertification and contribute to climate change adaptation and mitigation more extensively. This is due to the constraints on the use of indigenous and local knowledge and collective action, as well as economic and institutional barriers for SLM adoption (Banadda 2010 <sup>[[#fn:r1204|1204]]</sup> ; Cordingley et al. 2015 <sup>[[#fn:r1205|1205]]</sup> ; Lokonon and Mbaye 2018 <sup>[[#fn:r1206|1206]]</sup> ; Mulinge et al. 2016 <sup>[[#fn:r1207|1207]]</sup> ; Wildemeersch et al. 2015 <sup>[[#fn:r1208|1208]]</sup> ) (Section 3.1.4.2; 3.6.3). Sustainable development of drylands under these socio-economic and environmental (climate change, desertification) conditions will also depend on the ability of dryland agricultural households to diversify their livelihoods sources (Boserup 1965 <sup>[[#fn:r1209|1209]]</sup> ; Safriel and Adeel 2008 <sup>[[#fn:r1210|1210]]</sup> ).

<div id="section-3-6-2-2-socio-economic-responses-for-economic-diversification"></div>

<span id="socio-economic-responses-for-economic-diversification"></span>
==== 3.6.2.2 Socio-economic responses for economic diversification ====

<div id="section-3-6-2-2-socio-economic-responses-for-economic-diversification-block-1"></div>

'''Livelihood diversification''' through non-farm employment increases the resilience of rural households against desertification and extreme weather events by diversifying their income and consumption (high confidence). Moreover, it can provide the funds to invest into SLM (Belay et al. 2017 <sup>[[#fn:r1211|1211]]</sup> ; Bryan et al. 2009 <sup>[[#fn:r1212|1212]]</sup> ; Dumenu and Obeng 2016 <sup>[[#fn:r1213|1213]]</sup> ; Salik et al. 2017 <sup>[[#fn:r1214|1214]]</sup> ; Shiferaw et al. 2009 <sup>[[#fn:r1215|1215]]</sup> ). Access to non-agricultural employment is especially important for poorer pastoral households as their small herd sizes make them less resilient to drought (Fratkin 2013 <sup>[[#fn:r1216|1216]]</sup> ; Lybbert et al. 2004 <sup>[[#fn:r1217|1217]]</sup> ). However, access to alternative opportunities is limited in the rural areas of many developing countries, especially for women and marginalised groups who lack education and social networks (Reardon et al. 2008 <sup>[[#fn:r1218|1218]]</sup> ).

'''Migration''' is frequently used as an adaptation strategy to environmental change ( ''medium confidence'' ). Migration is a form of livelihood diversification and a potential response option to desertification and increasing risk to agricultural livelihoods under climate change (Walther et al. 2002 <sup>[[#fn:r1219|1219]]</sup> ). Migration can be short-term (e.g., seasonal) or long-term, internal within a country or international. There is ''medium evidence'' showing rural households responding to desertification and droughts through all forms of migration, for example: during the Dust Bowl in the USA in the 1930s (Hornbeck 2012 <sup>[[#fn:r1220|1220]]</sup> ); during droughts in Burkina Faso in the 2000s (Barbier et al. 2009 <sup>[[#fn:r1221|1221]]</sup> ); in Mexico in the 1990s (Nawrotzki et al. 2016 <sup>[[#fn:r1222|1222]]</sup> ); and by the Aymara people of the semi-arid Tarapacá region in Chile between 1820 and 1970, responding to declines in rainfall and growing demands for labour outside the region (Lima et al. 2016 <sup>[[#fn:r1223|1223]]</sup> ). There is ''robust evidence'' and ''high agreement'' showing that migration decisions are influenced by a complex set of different factors, with desertification and climate change playing relatively lesser roles (Liehr et al. 2016 <sup>[[#fn:r1224|1224]]</sup> ) (Section 3.4.2). Barrios et al. (2006) <sup>[[#fn:r1225|1225]]</sup> found that urbanisation in Sub-Saharan Africa was partially influenced by climatic factors during the 1950–2000 period, in parallel to liberalisation of internal restrictions on labour movements: each 1% reduction in rainfall was associated with a 0.45% increase in urbanisation. This migration favoured more industrially diverse urban areas in Sub-Saharan Africa (Henderson et al. 2017 <sup>[[#fn:r1226|1226]]</sup> ), because they offer more diverse employment opportunities and higher wages. Similar trends were also observed in Iran in response to water scarcity (Madani et al. 2016 <sup>[[#fn:r1227|1227]]</sup> ).

However, migration involves some initial investments. For this reason, reductions in agricultural incomes due to climate change or desertification have the potential to decrease out-migration among the poorest agricultural households, who become less able to afford migration (Cattaneo and Peri 2016 <sup>[[#fn:r1228|1228]]</sup> ), thus increasing social inequalities. There is ''medium evidence'' and high agreement that households with migrant worker members are more resilient against extreme weather events and environmental degradation compared to non-migrant households, who are more dependent on agricultural income (Liehr et al. 2016 <sup>[[#fn:r1229|1229]]</sup> ; Salik et al. 2017 <sup>[[#fn:r1230|1230]]</sup> ; Sikder and Higgins 2017 <sup>[[#fn:r1231|1231]]</sup> ). Remittances from migrant household members potentially contribute to SLM adoptions, however, substantial out-migration was also found to constrain the implementation of labour-intensive land management practices (Chen et al. 2014 <sup>[[#fn:r1232|1232]]</sup> ; Liu et al. 2016a <sup>[[#fn:r1233|1233]]</sup> ).

<span id="policy-responses"></span>
=== 3.6.3 Policy responses ===

<div id="section-3-6-3-policy-responses-block-1"></div>

The adoption of SLM practices depends on the compatibility of the technology with prevailing socio-economic and biophysical conditions (Sanz et al. 2017 <sup>[[#fn:r1798|1798]]</sup> ). Globally, it was shown that every USD invested into restoring degraded lands yields social returns, including both provisioning and non-provisioning ecosystem services, in the range of 3–6 USD over a 30-year period (Nkonya et al. 2016a <sup>[[#fn:r1234|1234]]</sup> ). A similar range of returns from land restoration activities was found in Central Asia (Mirzabaev et al. 2016 <sup>[[#fn:r1235|1235]]</sup> ), Ethiopia (Gebreselassie et al. 2016 <sup>[[#fn:r1236|1236]]</sup> ), India (Mythili and Goedecke 2016 <sup>[[#fn:r1237|1237]]</sup> ), Kenya (Mulinge et al. 2016 <sup>[[#fn:r1238|1238]]</sup> ), Niger (Moussa et al. 2016 <sup>[[#fn:r1239|1239]]</sup> ) and Senegal (Sow et al. 2016 <sup>[[#fn:r1240|1240]]</sup> ) ( ''medium confidence'' ). Despite these relatively high returns, there is ''robust evidence'' that the adoption of SLM practices remains low (Cordingley et al. 2015 <sup>[[#fn:r1241|1241]]</sup> ; Giger et al. 2015 <sup>[[#fn:r1242|1242]]</sup> ; Lokonon and Mbaye 2018 <sup>[[#fn:r1243|1243]]</sup> ). Part of the reason for these low adoption rates is that the major share of the returns from SLM are social benefits, namely in the form of non-provisioning ecosystem services (Nkonya et al. 2016a <sup>[[#fn:r1244|1244]]</sup> ). The adoption of SLM technologies does not always provide implementers with immediate private benefits (Schmidt et al. 2017 <sup>[[#fn:r1245|1245]]</sup> ). High initial investment costs, institutional and governance constraints and a lack of access to technologies and equipment may inhibit their adoption further (Giger et al. 2015 <sup>[[#fn:r1246|1246]]</sup> ; Sanz et al. 2017 <sup>[[#fn:r1247|1247]]</sup> ; Schmidt et al. 2017 <sup>[[#fn:r1248|1248]]</sup> ). However, not all SLM practices have high upfront costs. Analysing the World Overview of Conservation Approaches and Technologies (WOCAT) database, a globally acknowledged reference database for SLM, Giger et al. (2015) <sup>[[#fn:r1249|1249]]</sup> found that the upfront costs of SLM technologies ranged from about 20 USD to 5000 USD, with the median cost being around 500 USD. Many SLM technologies are profitable within 3 to 10 years ( ''medium confidence'' ) (Djanibekov and Khamzina 2016 <sup>[[#fn:r1250|1250]]</sup> ; Giger et al. 2015 <sup>[[#fn:r1251|1251]]</sup> ; Moussa et al. 2016 <sup>[[#fn:r1252|1252]]</sup> ; Sow et al. 2016 <sup>[[#fn:r1253|1253]]</sup> ). About 73% of 363 SLM technologies evaluated were reported to become profitable within three years, while 97% were profitable within 10 years (Giger et al. 2015 <sup>[[#fn:r1254|1254]]</sup> ). Similarly, it was shown that social returns from investments in restoring degraded lands will exceed their costs within six years in many settings across drylands (Nkonya et al. 2016a <sup>[[#fn:r1255|1255]]</sup> ). However, even with affordable upfront costs, market failures – in the form of lack of access to credit, input and output markets, and insecure land tenure (Section 3.1.3) – result in the lack of adoption of SLM technologies (Moussa et al. 2016 <sup>[[#fn:r1256|1256]]</sup> ). Payments for ecosystem services, subsidies for SLM, and encouragement of community collective action can lead to a higher level of adoption of SLM and land restoration activities ( ''medium confidence'' ) (Bouma and Wösten 2016 <sup>[[#fn:r1257|1257]]</sup> ; Lambin et al. 2014 <sup>[[#fn:r1258|1258]]</sup> ; Reed et al. 2015 <sup>[[#fn:r1259|1259]]</sup> ; Schiappacasse et al. 2012 <sup>[[#fn:r1260|1260]]</sup> ; van Zanten et al. 2014 <sup>[[#fn:r1261|1261]]</sup> ) (Section 3.6.3). Enabling the policy responses discussed in this section will contribute to overcoming these market failures.

Many socio-economic factors shaping individual responses to desertification typically operate at larger scales. Individual households and communities do not exercise control over these factors, such as land tenure insecurity, lack of property rights, lack of access to markets, availability of rural advisory services, and agricultural price distortions. These factors are shaped by national government policies and international markets. As is the case with socio-economic responses, policy responses are classified below in two ways: those which seek to combat desertification under changing climate; and those which seek to provide alternative livelihood sources through economic diversification. These options are mutually complementary and contribute to all the three hierarchical elements of the Land Degradation Neutrality (LDN) framework, namely, avoiding, reducing and reversing land degradation (Cowie et al. 2018 <sup>[[#fn:r1262|1262]]</sup> ; Orr et al. 2017 <sup>[[#fn:r1263|1263]]</sup> ) (Sections 4.8.5 and 7.4.5, and Table 7.2). An enabling policy environment is a critical element for the achievement of LDN (Chasek et al. 2019 <sup>[[#fn:r1264|1264]]</sup> ). Implementation of LDN policies can contribute to climate change adaptation and mitigation ( ''high confidence'' ) (Sections 3.6.1 and 3.7.2).

<div id="section-3-6-3-1-policy-responses-towards-combating-desertification-under-climate-change"></div>

<span id="policy-responses-towards-combating-desertification-under-climate-change"></span>
==== 3.6.3.1 Policy responses towards combating desertification under climate change ====

<div id="section-3-6-3-1-policy-responses-towards-combating-desertification-under-climate-change-block-1"></div>

Policy responses to combat desertification take numerous forms (Marques et al. 2016 <sup>[[#fn:r1265|1265]]</sup> ). Below we discuss major policy responses consistently highlighted in the literature in connection with SLM and climate change, because these response options were found to strengthen adaptation capacities and to contribute to climate change mitigation. They include improving market access, empowering women, expanding access to agricultural advisory services, strengthening land tenure security, payments for ecosystem services, decentralised natural resource management, investing into research and monitoring of desertification and dust storms, and investing into modern renewable energy sources.

'''Policies aiming at improving market access,''' that is the ability to access output and input markets at lower costs, help farmers and livestock producers earn more profit from their produce. Increased profits both motivate and enable them to invest more in SLM. Higher access to input, output and credit markets was consistently found as a major factor in the adoption of SLM practices in a wide number of settings across the drylands ( ''medium confidence'' ) (Aw-Hassan et al. 2016 <sup>[[#fn:r1266|1266]]</sup> ; Gebreselassie et al. 2016 <sup>[[#fn:r1267|1267]]</sup> ; Mythili and Goedecke 2016 <sup>[[#fn:r1268|1268]]</sup> ; Nkonya and Anderson 2015 <sup>[[#fn:r1269|1269]]</sup> ; Sow et al. 2016). Lack of access to credit limits adjustments and agricultural responses to the impacts of desertification under changing climate, with long-term consequences for the livelihoods and incomes, as was shown during the North American Dust Bowl of the 1930s (Hornbeck 2012 <sup>[[#fn:r1271|1271]]</sup> ). Government policies aimed at improving market access usually involve constructing and upgrading rural–urban transportation infrastructure and agricultural value chains, such as investments into construction of local markets, abattoirs and cold storage warehouses, as well as post-harvest processing facilities (McPeak et al. 2006). However, besides infrastructural constraints, providing improved access often involves relieving institutional constraints to market access (Little 2010 <sup>[[#fn:r1272|1272]]</sup> ), such as improved coordination of cross-border food safety and veterinary regulations (Ait Hou et al. 2015 <sup>[[#fn:r1273|1273]]</sup> ; Keiichiro et al. 2015 <sup>[[#fn:r1274|1274]]</sup> ; McPeak et al. 2006; Unnevehr 2015 <sup>[[#fn:r1275|1275]]</sup> ), and availability and access to market information systems (Bobojonov et al. 2016 <sup>[[#fn:r1276|1276]]</sup> ; Christy et al. 2014 <sup>[[#fn:r1277|1277]]</sup> ; Nakasone et al. 2014 <sup>[[#fn:r1278|1278]]</sup> ).

'''Women’s empowerment.''' A greater emphasis on understanding gender-specific differences over land use and land management practices as an entry point can make land restoration projects more successful ( ''medium confidence'' ) (Broeckhoven and Cliquet 2015 <sup>[[#fn:r1279|1279]]</sup> ; Carr and Thompson 2014 <sup>[[#fn:r1280|1280]]</sup> ; Catacutan and Villamor 2016 <sup>[[#fn:r1281|1281]]</sup> ; Dah-gbeto and Villamor 2016 <sup>[[#fn:r1282|1282]]</sup> ). In relation to representation and authority to make decisions in land management and governance, women’s participation remains lacking particularly in the dryland regions. Thus, ensuring women’s rights means accepting women as equal members of the community and citizens of the state (Nelson et al. 2015 <sup>[[#fn:r1283|1283]]</sup> ). This includes equitable access of women to resources (including extension services), networks, and markets. In areas where socio-cultural norms and practices devalue women and undermine their participation, actions for empowering women will require changes in customary norms, recognition of women’s (land) rights in government policies, and programmes to assure that their interests are better represented (Section 1.4.2 and Cross-Chapter Box 11 in Chapter 7). In addition, several novel concepts are recently applied for an in-depth understanding of gender in relation to science–policy interface. Among these are the concepts of intersectionality, that is, how social dimensions of identity and gender are bound up in systems of power and social institutions (Thompson-Hall et al. 2016 <sup>[[#fn:r1284|1284]]</sup> ), bounded rationality for gendered decision-making, related to incomplete information interacting with limits to human cognition leading to judgement errors or objectively poor decision making (Villamor and van Noordwijk 2016 <sup>[[#fn:r1285|1285]]</sup> ), anticipatory learning for preparing for possible contingencies and consideration of long-term alternatives (Dah-gbeto and Villamor 2016 <sup>[[#fn:r1286|1286]]</sup> ) and systematic leverage points for interventions that produce, mark, and entrench gender inequality within communities (Manlosa et al. 2018 <sup>[[#fn:r1287|1287]]</sup> ), which all aim to improve gender equality within agroecological landscapes through a systems approach.

'''Education and expanding access to agricultural services. ''' Providing access to information about SLM practices facilitates their adoption ( ''medium confidence'' ) (Kassie et al. 2015 <sup>[[#fn:r1288|1288]]</sup> ; Nkonya et al. 2015 <sup>[[#fn:r1289|1289]]</sup> ; Nyanga et al. 2016 <sup>[[#fn:r1291|1291]]</sup> ). Moreover, improving the knowledge of climate change, capacity building and development in rural areas can help strengthen climate change adaptive capacities (Berman et al. 2012 <sup>[[#fn:r1292|1292]]</sup> ; Chen et al. 2018 <sup>[[#fn:r1293|1293]]</sup> ; Descheemaeker et al. 2018 <sup>[[#fn:r1294|1294]]</sup> ; Popp et al. 2009 <sup>[[#fn:r1296|1296]]</sup> ; Tambo 2016 <sup>[[#fn:r1297|1297]]</sup> ; Yaro et al. 2015 <sup>[[#fn:r1298|1298]]</sup> ). Agricultural initiatives to improve the adaptive capacities of vulnerable populations were more successful when they were conducted through reorganised social institutions and improved communication, for example, in Mozambique (Osbahr et al. 2008 <sup>[[#fn:r1299|1299]]</sup> ). Improved communication and education could be facilitated by wider use of new information and communication technologies (ICTs) (Peters et al. 2015 <sup>[[#fn:r1300|1300]]</sup> ). Investments into education were associated with higher adoption of soil conservation measures, for example, in Tanzania (Tenge et al. 2004 <sup>[[#fn:r1301|1301]]</sup> ). Bryan et al. (2009) found that access to information was the prominent facilitator of climate change adaptation in Ethiopia. However, resource constraints of agricultural services, and disconnects between agricultural policy and climate policy can hinder the dissemination of climate-smart agricultural technologies (Morton 2017 <sup>[[#fn:r1302|1302]]</sup> ). Lack of knowledge was also found to be a significant barrier to implementation of soil rehabilitation programmes in the Mediterranean region (Reichardt 2010 <sup>[[#fn:r1303|1303]]</sup> ). Agricultural services will be able to facilitate SLM best when they also serve as platforms for sharing indigenous and local knowledge and farmer innovations (Mapfumo et al. 2016 <sup>[[#fn:r1304|1304]]</sup> ). Participatory research initiatives conducted jointly with farmers have higher chances of resulting in technology adoption (Bonney et al. 2016 <sup>[[#fn:r1305|1305]]</sup> ; Rusike et al. 2006 <sup>[[#fn:r1306|1306]]</sup> ; Vente et al. 2016). Moreover, rural advisory services are often more successful in disseminating technological innovations when they adopt commodity/value chain approaches, remain open to engagement in input supply, make use of new opportunities presented by ICTs, facilitate mutual learning between multiple stakeholders (Morton 2017 <sup>[[#fn:r1307|1307]]</sup> ), and organise science and SLM information in a location-specific manner for use in education and extension (Bestelmeyer et al. 2017 <sup>[[#fn:r1308|1308]]</sup> ).

'''Strengthening land tenure security.''' Strengthening land tenure security is a major factor contributing to the adoption of soil conservation measures in croplands ( ''high confidence'' ) (Bambio and Bouayad Agha 2018 <sup>[[#fn:r1309|1309]]</sup> ; Higgins et al. 2018 <sup>[[#fn:r1310|1310]]</sup> ; Holden and Ghebru 2016 <sup>[[#fn:r1311|1311]]</sup> ; Paltasingh 2018 <sup>[[#fn:r1312|1312]]</sup> ; Rao et al. 2016; Robinson et al. 2018 <sup>[[#fn:r1313|1313]]</sup> ), thus contributing to climate change adaptation and mitigation. Moreover, land tenure security can lead to more investment in trees (Deininger and Jin 2006 <sup>[[#fn:r1314|1314]]</sup> ; Etongo et al. 2015 <sup>[[#fn:r1315|1315]]</sup> ). Land tenure recognition policies were found to lead to higher agricultural productivity and incomes, although with inter-regional variations, requiring an improved understanding of overlapping formal and informal land tenure rights (Lawry et al. 2017 <sup>[[#fn:r1316|1316]]</sup> ). For example, secure land tenure increased investments into SLM practices in Ghana, but without affecting farm productivity (Abdulai et al. 2011 <sup>[[#fn:r1317|1317]]</sup> ). Secure land tenure, especially for communally managed lands, helps reduce arbitrary appropriations of land for large-scale commercial farms (Aha and Ayitey 2017; Baumgartner 2017 <sup>[[#fn:r1318|1318]]</sup> ; Dell’Angelo et al. 2017 <sup>[[#fn:r1319|1319]]</sup> ). In contrast, privatisation of rangeland tenures in Botswana and Kenya led to the loss of communal grazing lands and actually increased rangeland degradation (Basupi et al. 2017 <sup>[[#fn:r1320|1320]]</sup> ; Kihiu 2016 <sup>[[#fn:r1321|1321]]</sup> ) as pastoralists needed to graze livestock on now smaller communal pastures. Since food insecurity in drylands is strongly affected by climate risks, there is ''robust evidence'' and ''high agreement'' that resilience to climate risks is higher with flexible tenure for allowing mobility for pastoralist communities, and not fragmenting their areas of movement (Behnke 1994 <sup>[[#fn:r1323|1323]]</sup> ; Holden and Ghebru 2016 <sup>[[#fn:r1324|1324]]</sup> ; Liao et al. 2017 <sup>[[#fn:r1325|1325]]</sup> ; Turner et al. 2016 <sup>[[#fn:r1326|1326]]</sup> ; Wario et al. 2016 <sup>[[#fn:r1327|1327]]</sup> ). More research is needed on the optimal tenure mix, including low-cost land certification, redistribution reforms, market-assisted reforms and gender-responsive reforms, as well as collective forms of land tenure such as communal land tenure and cooperative land tenure (see Section 7.6.5 for a broader discussion of land tenure security under climate change).

'''Payment for ecosystem services (PES)''' provides incentives for land restoration and SLM ( ''medium confidence'' ) (Lambin et al. 2014 <sup>[[#fn:r1328|1328]]</sup> ; Li et al. 2018; Reed et al. 2015 <sup>[[#fn:r1329|1329]]</sup> ; Schiappacasse et al. 2012 <sup>[[#fn:r1330|1330]]</sup> ). Several studies illustrate that the social costs of desertification are larger than its private cost (Costanza et al. 2014 <sup>[[#fn:r1331|1331]]</sup> ; Nkonya et al. 2016a <sup>[[#fn:r1332|1332]]</sup> ). Therefore, although SLM can generate public goods in the form of provisioning ecosystem services, individual land custodians underinvest in SLM as they are unable to reap these benefits fully. Payment for ecosystem services provides a mechanism through which some of these benefits can be transferred to land users, thereby stimulating further investment in SLM. The effectiveness of PES schemes depends on land tenure security and appropriate design, taking into account specific local conditions (Börner et al. 2017 <sup>[[#fn:r1333|1333]]</sup> ). However, PES has not worked well in countries with fragile institutions (Karsenty and Ongolo 2012 <sup>[[#fn:r1334|1334]]</sup> ). Equity and justice in distributing the payments for ecosystem services were found to be key for the success of the PES programmes in Yunnan, China (He and Sikor 2015). Yet, when reviewing the performance of PES programmes in the tropics, Calvet-Mir et al. (2015), found that they are generally effective in terms of environmental outcomes, despite being sometimes unfair in terms of payment distribution. It is suggested that the implementation of PES will be improved through decentralised approaches giving local communities a larger role in the decision-making process (He and Lang 2015).

'''Empowering local communities for decentralised natural resource management.''' Local institutions often play a vital role in implementing SLM initiatives and climate change adaptation measures ( ''high confidence'' ) (Gibson et al. 2005 <sup>[[#fn:r1335|1335]]</sup> ; Smucker et al. 2015 <sup>[[#fn:r1336|1336]]</sup> ). Pastoralists involved in community-based natural resource management in Mongolia had greater capacity to adapt to extreme winter frosts, resulting in less damage to their livestock (Fernandez-Gimenez et al. 2015 <sup>[[#fn:r1337|1337]]</sup> ). Decreasing the power and role of traditional community institutions, due to top-down public policies, resulted in lower success rates in community-based programmes focused on rangeland management in Dirre, Ethiopia (Abdu and Robinson 2017 <sup>[[#fn:r1338|1338]]</sup> ). Decentralised governance was found to lead to improved management in forested landscapes (Dressler et al. 2010 <sup>[[#fn:r1339|1339]]</sup> ; Ostrom and Nagendra 2006 <sup>[[#fn:r1340|1340]]</sup> ). However, there are also cases when local elites were placed in control and this decentralised natural resource management negatively impacted the livelihoods of the poorer and marginalised community members due to reduced access to natural resources (Andersson and Ostrom 2008 <sup>[[#fn:r1341|1341]]</sup> ; Cullman 2015 <sup>[[#fn:r1343|1343]]</sup> ; Dressler et al. 2010 <sup>[[#fn:r1344|1344]]</sup> ).

The success of decentralised natural resource management initiatives depends on increased participation and empowerment of a diverse set of community members, not only local leaders and elites, in the design and management of local resource management institutions (Kadirbeyoglu and Özertan 2015 <sup>[[#fn:r1345|1345]]</sup> ; Umutoni et al. 2016 <sup>[[#fn:r1346|1346]]</sup> ), while considering the interactions between actors and institutions at different levels of governance (Andersson and Ostrom 2008 <sup>[[#fn:r1347|1347]]</sup> ; Carlisle and Gruby 2017 <sup>[[#fn:r1349|1349]]</sup> ; McCord et al. 2017 <sup>[[#fn:r1351|1351]]</sup> ). An example of such programmes where local communities played a major role in land restoration and rehabilitation activities is the cooperative project on The National Afforestation and Erosion Control Mobilization Action Plan in Turkey, initiated by the Turkish Ministry of Agriculture and Forestry (Çalişkan and Boydak 2017 <sup>[[#fn:r1352|1352]]</sup> ), with the investment of 1.8 billion USD between 2008 and 2012. The project mobilised local communities in cooperation with public institutions, municipalities, and non-governmental organisations, to implement afforestation, rehabilitation and erosion control measures, resulting in the afforestation and reforestation of 1.5 Mha (Yurtoglu 2015 <sup>[[#fn:r1353|1353]]</sup> ). Moreover, some 1.75 Mha of degraded forest and 37,880 ha of degraded rangelands were rehabilitated. Finally, the project provided employment opportunities for 300,000 rural residents for six months every year, combining land restoration and rehabilitation activities with measures to promote socio-economic development in rural areas (Çalişkan and Boydak 2017 <sup>[[#fn:r1354|1354]]</sup> ).

'''Investing in research and development.''' Desertification has received substantial research attention over recent decades (Turner et al. 2007 <sup>[[#fn:r1355|1355]]</sup> ). There is also a growing research interest on climate change adaptation and mitigation interventions that help address desertification (Grainger 2009 <sup>[[#fn:r1356|1356]]</sup> ). Agricultural research on SLM practices has generated a significant number of new innovations and technologies that increase crop yields without degrading the land, while contributing to climate change adaptation and mitigation (Section 3.6.1). There is ''robust evidence'' that such technologies help improve the food security of smallholder dryland farming households (Harris and Orr 2014 <sup>[[#fn:r1357|1357]]</sup> ) (Section 6.3.5). Strengthening research on desertification is of high importance not only to meet SDGs but also to manage ecosystems effectively, based on solid scientific knowledge. More investment in research institutes and training the younger generation of researchers is needed for addressing the combined challenges of desertification and climate change (Akhtar-Schuster et al. 2011 <sup>[[#fn:r1358|1358]]</sup> ; Verstraete et al. 2011 <sup>[[#fn:r1359|1359]]</sup> ). This includes improved knowledge management systems that allow stakeholders to work in a coordinated manner by enhancing timely, targeted and contextualised information sharing (Chasek et al. 2011 <sup>[[#fn:r1360|1360]]</sup> ). Knowledge and flow of knowledge on desertification is currently highly fragmented, constraining the effectiveness of those engaged in assessing and monitoring the phenomenon at various levels (Reed et al. 2011 <sup>[[#fn:r1361|1361]]</sup> ). Improved knowledge and data exchange and sharing increase the effectiveness of efforts to address desertification ( ''high confidence'' ).

'''Developing modern renewable energy sources.''' Transitioning to renewable energy resources contributes to reducing desertification by lowering reliance on traditional biomass in dryland regions ( ''medium confidence'' ). This can also have socioeconomic and health benefits, especially for women and children ( ''high confidence'' ). Populations in most developing countries continue to rely on traditional biomass, including fuelwood, crop straws and livestock manure, for a major share of their energy needs, with the highest dependence in Sub-Saharan Africa (Amugune et al. 2017 <sup>[[#fn:r1363|1363]]</sup> ; IEA 2013). Use of biomass for energy, mostly fuelwood (especially as charcoal), was associated with deforestation in some dryland areas (Iiyama et al. 2014 <sup>[[#fn:r1364|1364]]</sup> ; Mekuria et al. 2018 <sup>[[#fn:r1365|1365]]</sup> ; Neufeldt et al. 2015 <sup>[[#fn:r1366|1366]]</sup> ; Zulu 2010 <sup>[[#fn:r1367|1367]]</sup> ), while in some other areas there was no link between fuelwood collection and deforestation (Simon and Peterson 2018 <sup>[[#fn:r1368|1368]]</sup> ; Swemmer et al. 2018 <sup>[[#fn:r1369|1369]]</sup> ; Twine and Holdo 2016 <sup>[[#fn:r1370|1370]]</sup> ). Moreover, the use of traditional biomass as a source of energy was found to have negative health effects through indoor air pollution (de la Sota et al. 2018 <sup>[[#fn:r1371|1371]]</sup> ; Lim and Seow 2012), while also being associated with lower female labour force participation (Burke and Dundas 2015 <sup>[[#fn:r1372|1372]]</sup> ). Jiang et al. (2014) indicated that providing improved access to alternative energy sources such as solar energy and biogas could help reduce the use of fuelwood in south-western China, thus alleviating the spread of rocky desertification. The conversion of degraded lands into cultivation of biofuel crops will affect soil carbon dynamics (Albanito et al. 2016 <sup>[[#fn:r1374|1374]]</sup> ; Nair et al. 2011 <sup>[[#fn:r1375|1375]]</sup> ) (Cross-Chapter Box 7 in Chapter 6). The use of biogas slurry as soil amendment or fertiliser can increase soil carbon (Galvez et al. 2012; Negash et al. 2017 <sup>[[#fn:r1376|1376]]</sup> ). Large-scale installation of wind and solar farms in the Sahara Desert was projected to create a positive climate feedback through increased surface friction and reduced albedo, doubling precipitation over the neighbouring Sahel region with resulting increases in vegetation (Li et al. 2018). Transition to renewable energy sources in high-income countries in dryland areas primarily contributes to reducing GHG emissions and mitigating climate change, with some other co-benefits such as diversification of energy sources (Bang 2010 <sup>[[#fn:r1377|1377]]</sup> ), while the impacts on desertification are less evident. The use of renewable energy has been proposed as an important mitigation option in dryland areas as well (El-Fadel et al. 2003 <sup>[[#fn:r1378|1378]]</sup> ). Transitions to renewable energy are being promoted by governments across drylands (Cancino-Solórzano et al. 2016 <sup>[[#fn:r1379|1379]]</sup> ; Hong et al. 2013 <sup>[[#fn:r1380|1380]]</sup> ; Sen and Ganguly 2017) including in fossil-fuel rich countries (Farnoosh et al. 2014 <sup>[[#fn:r1381|1381]]</sup> ; Dehkordi et al. 2017; Stambouli et al. 2012 <sup>[[#fn:r1382|1382]]</sup> ; Vidadili et al. 2017 <sup>[[#fn:r1383|1383]]</sup> ), despite important social, political and technical barriers to expanding renewable energy production (Afsharzade et al. 2016; Baker et al. 2014 <sup>[[#fn:r1384|1384]]</sup> ; Elum and Momodu 2017 <sup>[[#fn:r1385|1385]]</sup> ; Karatayev et al. 2016 <sup>[[#fn:r1386|1386]]</sup> ). Improving social awareness about the benefits of transitioning to renewable energy resources, and access to hydro-energy, solar and wind energy contributes to their improved adoption (Aliyu et al. 2017 <sup>[[#fn:r1387|1387]]</sup> ; Katikiro 2016).

'''Developing and strengthening climate services relevant for desertification.''' Climate services provide climate, drought and desertification-related information in a way that assists decision-making by individuals and organisations. Monitoring desertification, and integrating biogeophysical (climate, soil, ecological factors, biodiversity) and socio-economic (use of natural resources by local population) issues provide a basis for better vulnerability prediction and assessment (OSS, 2012; Vogt et al. 2011 <sup>[[#fn:r1388|1388]]</sup> ). Examples of relevant services include: drought monitoring and early warning systems, often implemented by national climate and meteorological services but also encompassing regional and global systems (Pozzi et al. 2013 <sup>[[#fn:r1389|1389]]</sup> ); and the Sand and Dust Storm Warning Advisory and Assessment System (SDS-WAS), created by WMO in 2007, in partnership with the World Health Organization (WHO) and the United Nations Environment Program (UNEP). Currently, there is also a lack of ecological monitoring in arid and semi-arid regions to study surface winds, dust and sand storms, and their impacts on ecosystems and human health (Bergametti et al. 2018 <sup>[[#fn:r1390|1390]]</sup> ; Marticorena et al. 2010 <sup>[[#fn:r1391|1391]]</sup> ). Reliable and timely climate services, relevant to desertification, can aid the development of appropriate adaptation and mitigation options, reducing the impact of desertification under changing climate on human and natural systems ( ''high confidence'' ) (Beegum et al. 2016 <sup>[[#fn:r1392|1392]]</sup> ; Beegum et al. 2018; Cornet 2012 <sup>[[#fn:r1393|1393]]</sup> ; Haase et al. 2018 <sup>[[#fn:r1395|1395]]</sup> ; Sergeant et al. 2012 <sup>[[#fn:r1396|1396]]</sup> ).

<div id="section-3-6-3-2-policy-responses-supporting-economic-diversification"></div>

<span id="policy-responses-supporting-economic-diversification"></span>
==== 3.6.3.2 Policy responses supporting economic diversification ====

<div id="section-3-6-3-2-policy-responses-supporting-economic-diversification-block-1"></div>

Despite policy responses for combating desertification, other factors will put strong pressures on the land, including climate change and growing food demands, as well as the need to reduce poverty and strengthen food security (Cherlet et al. 2018 <sup>[[#fn:r1397|1397]]</sup> ) (Sections 6.1.4 and 7.2.2). Sustainable development of drylands and their resilience to combined challenges of desertification and climate change will thus also depend on the ability of governments to promote policies for economic diversification within agriculture and in non-agricultural sectors in order make dryland areas less vulnerable to desertification and climate change.

'''Investing into irrigation.''' Investments into expanding irrigation in dryland areas can help increase the resilience of agricultural production to climate change, improve labour productivity and boost production and income revenue from agriculture and livestock sectors (Geerts and Raes 2009 <sup>[[#fn:r1399|1399]]</sup> ; Olayide et al. 2016 <sup>[[#fn:r1400|1400]]</sup> ; Oweis and Hachum 2006 <sup>[[#fn:r1401|1401]]</sup> ). This is particularly true for Sub-Saharan Africa, where currently only 6% of the cultivated areas are irrigated (Nkonya et al. 2016b <sup>[[#fn:r1402|1402]]</sup> ). While renewable groundwater resources could help increase the share of irrigated land to 20.5–48.6% of croplands in the region (Altchenko and Villholth 2015 <sup>[[#fn:r1403|1403]]</sup> ). On the other hand, over-extraction of groundwaters, mainly for irrigating crops, is becoming an important environmental problem in many dryland areas (Cherlet et al. 2018 <sup>[[#fn:r1404|1404]]</sup> ), requiring careful design and planning of irrigation expansion schemes and use of water-efficient irrigation methods (Bjornlund et al. 2017 <sup>[[#fn:r1405|1405]]</sup> ; Woodhouse et al. 2017 <sup>[[#fn:r1406|1406]]</sup> ). For example, in Saudi Arabia, improving the efficiency of water management, for example through the development of aquifers, water recycling and rainwater harvesting, is part of a suite of policy actions to combat desertification (Bazza, et al. 2018 <sup>[[#fn:r1407|1407]]</sup> ; Kingdom of Saudi Arabia 2016 <sup>[[#fn:r1408|1408]]</sup> ). The expansion of irrigation to riverine areas, crucial for dry season grazing of livestock, needs to consider the income from pastoral activities, which is not always lower than income from irrigated crop production (Behnke and Kerven 2013 <sup>[[#fn:r1409|1409]]</sup> ). Irrigation development could be combined with the deployment of clean-energy technologies in economically viable ways (Chandel et al. 2015 <sup>[[#fn:r1410|1410]]</sup> ). For example, solar-powered drip irrigation was found to increase household agricultural incomes in Benin (Burney et al. 2010 <sup>[[#fn:r1411|1411]]</sup> ). The sustainability of irrigation schemes based on solar-powered extraction of groundwaters depends on measures to avoid over-abstraction of groundwater resources and associated negative environmental impacts (Closas and Rap 2017 <sup>[[#fn:r1412|1412]]</sup> ).

'''Expanding agricultural commercialisation.''' Faster poverty rate reduction and economic growth enhancement is realised when countries transition into the production of non-staple, high-value commodities and manage to build a robust agro-industry sector (Barrett et al. 2017 <sup>[[#fn:r1413|1413]]</sup> ). Ogutu and Qaim (2019) found that agricultural commercialisation increased incomes and decreased multidimensional poverty in Kenya. Similar findings were earlier reported by Muriithi and Matz (2015) for commercialisation of vegetables in Kenya. Commercialisation of rice production was found to have increased smallholder welfare in Nigeria (Awotide et al. 2016 <sup>[[#fn:r1414|1414]]</sup> ). Agricultural commercialisation contributed to improved household food security in Malawi, Tanzania and Uganda (Carletto et al. 2017 <sup>[[#fn:r1415|1415]]</sup> ). However, such a transition did not improve farmers’ livelihoods in all cases (Reardon et al. 2009). High-value cash crop/animal production can be bolstered by wide-scale use of technologies, for example, mechanisation, application of inorganic fertilisers, crop protection and animal health products. Market oriented crop/animal production facilitates social and economic progress, with labour increasingly shifting out of agriculture into non-agricultural sectors (Cour 2001). Modernised farming, improved access to inputs, credit and technologies enhances competitiveness in local and international markets (Reardon et al. 2009 <sup>[[#fn:r1417|1417]]</sup> ).

'''Facilitating structural transformations''' in rural economies implies that the development of non-agricultural sectors encourages the movement of labour from land-based livelihoods, vulnerable to desertification and climate change, to non-agricultural activities (Haggblade et al. 2010 <sup>[[#fn:r1420|1420]]</sup> ). The movement of labour from agriculture to non-agricultural sectors is determined by relative labour productivities in these sectors (Shiferaw and Djido 2016 <sup>[[#fn:r1421|1421]]</sup> ). Given already high underemployment in the farm sector, increasing labour productivity in the non-farm sector was found as the main driver of labour movements from farm sector to non-farm sector (Shiferaw and Djido 2016 <sup>[[#fn:r1422|1422]]</sup> ). More investments into education can facilitate this process (Headey et al. 2014 <sup>[[#fn:r1423|1423]]</sup> ). However, in some contexts, such as pastoralist communities in Xinjiang, China, income diversification was not found to improve the welfare of pastoral households (Liao et al. 2015 <sup>[[#fn:r1424|1424]]</sup> ). Economic transformations also occur through urbanisation, involving the shift of labour from rural areas into gainful employment in urban areas (Jedwab and Vollrath 2015 <sup>[[#fn:r1425|1425]]</sup> ). The majority of world population will be living in urban centres in the 21st century and this will require innovative means of agricultural production with minimum ecological footprint and less dependence on fossil fuels (Revi and Rosenzweig 2013 <sup>[[#fn:r1426|1426]]</sup> ), while addressing the demand of cities (see Section 4.9.1 for discussion on urban green infrastructure). Although there is some evidence of urbanisation leading to the loss of indigenous and local ecological knowledge, however, indigenous and local knowledge systems are constantly evolving, and are also being integrated into urban environments (Júnior et al. 2016 <sup>[[#fn:r1427|1427]]</sup> ; Reyes-García et al. 2013 <sup>[[#fn:r1429|1429]]</sup> ; van Andel and Carvalheiro 2013 <sup>[[#fn:r1430|1430]]</sup> ). Urban areas are attracting an increasing number of rural residents across the developing world (Angel et al. 2011 <sup>[[#fn:r1431|1431]]</sup> ; Cour 2001 <sup>[[#fn:r1432|1432]]</sup> ; Dahiya 2012 <sup>[[#fn:r1433|1433]]</sup> ). Urban development contributes to expedited agricultural commercialisation by providing market outlet for cash crops, high-value crops, and livestock products. At the same time, urbanisation also poses numerous challenges in the form of rapid urban sprawl and pressures on infrastructure and public services, unemployment and associated social risks, which have considerable implications on climate change adaptive capacities (Bulkeley 2013 <sup>[[#fn:r1434|1434]]</sup> ; Garschagen and Romero-Lankao 2015 <sup>[[#fn:r1435|1435]]</sup> ).

<div id="section-3-6-3-2-policy-responses-supporting-economic-diversification-block-2" class="box"></div>

<span id="ccb5-policy-responses-to-drought"></span>