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

== 3.7 Hotspots and case studies ==

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The challenges of desertification and climate change in dryland areas across the world often have very location-specific characteristics. The five case studies in this section present rich experiences and lessons learnt on: (i) soil erosion, (ii) afforestation and reforestation through ‘green walls’, (iii) invasive plant species, (iv) oases in hyper-arid areas, and (v) integrated watershed management. Although it is impossible to cover all hotspots of desertification and on-the-ground actions from all dryland areas, these case studies present a more focused assessment of these five issues, which emerged as salient in the group discussions and several rounds of review of this chapter. The choice of these case studies was also motivated by the desire to capture a wide diversity of dryland settings.

<span id="climate-change-and-soil-erosion"></span>
=== 3.7.1 Climate change and soil erosion ===

<div id="section-3-7-1-1-soil-erosion-under-changing-climate-in-drylands"></div>

<span id="soil-erosion-under-changing-climate-in-drylands"></span>
==== 3.7.1.1 Soil erosion under changing climate in drylands ====

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Soil erosion is a major form of desertification occurring in varying degrees in all dryland areas across the world (Section 3.2), with negative effects on dryland ecosystems (Section 3.4). Climate change is projected to increase soil erosion potential in some dryland areas through more frequent heavy rainfall events and rainfall variability (see Section 3.5.2 for a more detailed assessment) (Achite and Ouillon 2007 <sup>[[#fn:r1500|1500]]</sup> ; Megnounif and Ghenim 2016 <sup>[[#fn:r1501|1501]]</sup> ; Vachtman et al. 2013 <sup>[[#fn:r1502|1502]]</sup> ; Zhang and Nearing 2005 <sup>[[#fn:r1503|1503]]</sup> ). There are numerous soil conservation measures that can help reduce soil erosion (Section 3.6.1). Such soil management measures include afforestation and reforestation activities, rehabilitation of degraded forests, erosion control measures, prevention of overgrazing, diversification of crop rotations, and improvement in irrigation techniques, especially in sloping areas (Anache et al. 2018 <sup>[[#fn:r1504|1504]]</sup> ; ÇEMGM 2017; Li and Fang 2016; Poesen 2018 <sup>[[#fn:r1505|1505]]</sup> ; Ziadat and Taimeh 2013 <sup>[[#fn:r1506|1506]]</sup> ). Effective measures for soil conservation can also use spatial patterns of plant cover to reduce sediment connectivity, and the relationships between hillslopes and sediment transfer in eroded channels (García-Ruiz et al. 2017 <sup>[[#fn:r1507|1507]]</sup> ). The following three examples present lessons learnt from the soil erosion problems and measures to address them in different settings of Chile, Turkey and the Central Asian countries.

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<span id="no-till-practices-for-reducing-soil-erosion-in-central-chile"></span>
==== 3.7.1.2 No-till practices for reducing soil erosion in central Chile ====

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Soil erosion by water is an important problem in Chile. National assessments conducted in 1979, which examined 46% of the continental surface of the country, concluded that very high levels of soil erosion affected 36% of the territory. The degree of soil erosion increases from south to north. The leading locations in Chile are the region of Coquimbo with 84% of eroded soils (Lat. 29°S, semi-arid climate), the region of Valparaíso with 57% of eroded soils (Lat. 33°S, Mediterranean climate) and the region of O’Higgins with 37% of eroded soils (Lat. 34°S, Mediterranean climate). The most important drivers of soil erosion are soil, slope, climate erosivity (i.e., precipitation, intensity, duration and frequency) due to a highly concentrated rainy season, and vegetation structure and cover. In the region of Coquimbo, goat and sheep overgrazing have aggravated the situation (CIREN 2010 <sup>[[#fn:r1508|1508]]</sup> ). Erosion rates reach up to 100 t ha <sup>–1</sup> annually, having increased substantially over the last 50 years (Ellies 2000). About 10.4% of central Chile exhibits high erosion rates (greater than 1.1 t ha <sup>–1</sup> annually) (Bonilla et al. 2010 <sup>[[#fn:r1509|1509]]</sup> ).

Over the last few decades there has been an increasing interest in the development of no-till (also called zero tillage) technologies to minimise soil disturbance, reduce the combustion of fossil fuels and increase soil organic matter. No-till, in conjunction with the adoption of strategic cover crops, has positively impacted soil biology with increases in soil organic matter. Early evaluations by Crovetto, (1998) showed that no-till application (after seven years) had doubled the biological activity indicators compared to traditional farming and even surpassed those found in pasture (grown for the previous 15 years). Besides erosion control, additional benefits are an increase of water-holding capacity and reduction in bulk density. Currently, the above no-till farm experiment has lasted for 40 years and continues to report benefits to soil health and sustainable production (Reicosky and Crovetto 2014 <sup>[[#fn:r1510|1510]]</sup> ). The influence of this iconic farm has resulted in the adoption of soil conservation practices – and especially no-till – in dryland areas of the Mediterranean climate region of central Chile (Martínez et al. 2011 <sup>[[#fn:r1511|1511]]</sup> ). Currently, it has been estimated that the area under no-till farming in Chile varies between 0.13 and 0.2 Mha (Acevedo and Silva 2003 <sup>[[#fn:r1512|1512]]</sup> ).

<div id="section-3-7-1-3-combating-wind-erosion-and-deflation-in-turkey-the-greening-desert-of-karapinar"></div>

<span id="combating-wind-erosion-and-deflation-in-turkey-the-greening-desert-of-karapınar"></span>
==== 3.7.1.3 Combating wind erosion and deflation in Turkey: The greening desert of Karapınar ====

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In Turkey, the amount of sediment recently released through erosion into seas was estimated to be 168 Mt yr <sup>-1</sup> , which is considerably lower than the 500 Mt yr <sup>–1</sup> that was estimated to be lost in the 1970s. The decrease in erosion rates is attributed to an increase in spatial extent of forests, rehabilitation of degraded forests, erosion control, prevention of overgrazing, and improvement in irrigation technologies. Soil conservation measures conducted in the Karapınar district, Turkey, exemplify these activities. The district is characterised by a semi-arid climate and annual average precipitation of 250–300 mm (Türkeş 2003 <sup>[[#fn:r1513|1513]]</sup> ; Türkeş and Tatlı 2011 <sup>[[#fn:r1514|1514]]</sup> ). In areas where vegetation was overgrazed or inappropriately tilled, the surface soil horizon was removed through erosion processes resulting in the creation of large drifting dunes that threatened settlements around Karapınar (Groneman 1968 <sup>[[#fn:r1515|1515]]</sup> ). Such dune movement had begun to affect the Karapınar settlement in 1956 (Kantarcı et al. 2011 <sup>[[#fn:r1516|1516]]</sup> ). Consequently, by the early 1960s, Karapınar town and nearby villages were confronted with the danger of abandonment due to out-migration in the early 1960s (Figure 3.11(1)). The reasons for increasing wind erosion in the Karapınar district can be summarised as follows: sandy material was mobilised following drying of the lake; hot and semi-arid climate conditions; overgrazing and use of pasture plants for fuel; excessive tillage; and strong prevailing winds.

<div id="section-3-7-1-3-combating-wind-erosion-and-deflation-in-turkey-the-greening-desert-of-karapinar-block-2"></div>

<span id="figure-3.11a"></span>
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'''Figure 3.11a'''

<span id="a-general-view-of-a-nearby-village-of-karapınar-town-in-the-early-1960s-çarkaci-1999."></span>
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'''(1) A general view of a nearby village of Karapınar town in the early 1960s (Çarkaci 1999).'''
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[[File:94b8ddda39bc1f5c6cec209097079501 Figure-3.11a.png]]

(1) A general view of a nearby village of Karapınar town in the early 1960s (Çarkaci 1999) <sup>[[#fn:r1802|1802]]</sup> .

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<span id="figure-3.11b"></span>
<!-- START IMG -->
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'''Figure 3.11b'''

<span id="a-view-of-the-karapınar-wind-erosion-area-in-2013-photo-murat-türkeş-17-june-2019."></span>
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'''(2)A view of the Karapınar wind erosion area in 2013 (Photo: Murat Türkeş, 17 June 2019).'''
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[[File:332259ecd3c1f439336fa3143309524a Figure-3.11b-1024x683.jpg]]

(2)A view of the Karapınar wind erosion area in 2013 (Photo: Murat Türkeş <sup>[[#fn:r1803|1803]]</sup> , 17 June 2019).

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<span id="figure-3.11c"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.11c'''

<span id="construction-of-cane-screens-in-the-early-1960s-in-order-to-decrease-wind-speed-and-prevent-movement-of-the-sand-accumulations-and-dunes-this-was-one-of-the-physical-measures-during-the-prevention-and-mitigation-period-çarkaci-1999."></span>
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'''(3) Construction of cane screens in the early 1960s in order to decrease wind speed and prevent movement of the sand accumulations and dunes; this was one of the physical measures during the prevention and mitigation period (Çarkaci 1999).'''

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[[File:ef030280cb1db9bd660d1c4f6d826f54 Figure-3.11c.png]]

(3) Construction of cane screens in the early 1960s in order to decrease wind speed and prevent movement of the sand accumulations and dunes; this was one of the physical measures during the prevention and mitigation period (Çarkaci 1999 <sup>[[#fn:r1804|1804]]</sup> ).

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<div id="section-3-7-1-3-combating-wind-erosion-and-deflation-in-turkey-the-greening-desert-of-karapinar-block-5"></div>

<span id="figure-3.11d"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.11d'''

<span id="a-view-of-mixed-vegetation-which-now-covers-most-of-the-karapınar-wind-erosion-area-in-2013-the-main-tree-species-of-which-were-selected-for-afforestation-with-respect-to-their-resistance-to-the-arid-continental-climate-conditions-along-with-a-warmhot-temperature-regime-over-the-district-photo-murat-türkeş-17-june-2013."></span>
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'''(4) A view of mixed vegetation, which now covers most of the Karapınar wind erosion area in 2013, the main tree species of which were selected for afforestation with respect to their resistance to the arid continental climate conditions along with a warm/hot temperature regime over the district (Photo: Murat Türkeş, 17 June 2013).'''
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[[File:4bca06773595a7299c60fb68e7be04f5 Figure-3.11d.png]]

(4) A view of mixed vegetation, which now covers most of the Karapınar wind erosion area in 2013, the main tree species of which were selected for afforestation with respect to their resistance to the arid continental climate conditions along with a warm/hot temperature regime over the district (Photo: Murat Türkeş <sup>[[#fn:r1805|1805]]</sup> , 17 June 2013).

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Restoration and mitigation strategies were initiated in 1959, and today 4300 ha of land have been restored (Akay and Yildirim 2010 <sup>[[#fn:r1517|1517]]</sup> ) (Figure 3.11 (2)), using specific measures: (i) physical measures: construction of cane screens to decrease wind speed and prevent sand movement (Figure 3.11(3)); (ii) restoration of cover: increasing grass cover between screens using seeds collected from local pastures or the cultivation of rye ( ''Secale'' sp.) and wheat grass ( ''Agropyron elongatum'' ) that are known to grow in arid and hot conditions; and (iii) afforestation: saplings obtained from nursery gardens were planted and grown between these screens. Main tree species selected were oleaster ( ''Eleagnus'' sp.), acacia ( ''Robinia pseudeaccacia'' ), ash ( ''Fraxinus'' sp.), elm ( ''Ulmus'' sp.) and maple (Acer sp.) (Figure 3.11 (4)). Economic growth occurred after controlling erosion and new tree nurseries have been established with modern irrigation. Potential negative consequences through the excessive use of water can be mitigated through engagement with local stakeholders and transdisciplinary learning processes, as well as by restoring the traditional land uses in the semi-arid Konya closed basin (Akça et al. 2016 <sup>[[#fn:r1518|1518]]</sup> ).

<div id="section-3-7-1-4-soil-erosion-in-central-asia-under-changing-climate"></div>

<span id="soil-erosion-in-central-asia-under-changing-climate"></span>

==== 3.7.1.4 Soil erosion in Central Asia under changing climate ====

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Soil erosion is widely acknowledged to be a major form of degradation of Central Asian drylands, affecting a considerable share of croplands and rangelands. However, up-to-date information on the actual extent of eroded soils at the regional or country level is not available. The estimates compiled by Pender et al. (2009), based on the Central Asian Countries Initiative for Land Management (CACILM), indicate that about 0.8 Mha of the irrigated croplands were subject to high degree of soil erosion in Uzbekistan. In Turkmenistan, soil erosion was indicated to be occurring in about 0.7 Mha of irrigated land. In Kyrgyzstan, out of 1 Mha of irrigated land in the foothill zones, 0.76 Mha were subject to soil erosion by water, leading to losses in crop yields of 20–60% in these eroded soils. About 0.65 Mha of arable land were prone to soil erosion by wind (Mavlyanova et al. 2017 <sup>[[#fn:r1519|1519]]</sup> ). Soil erosion is widespread in rainfed and irrigated areas in Kazakhstan (Saparov 2014). About 5 Mha of rainfed croplands were subject to high levels of soil erosion (Pender et al. 2009 <sup>[[#fn:r1520|1520]]</sup> ). Soil erosion by water was indicated to be a major concern in sloping areas in Tajikistan (Pender et al. 2009 <sup>[[#fn:r1521|1521]]</sup> ).

The major causes of soil erosion in Central Asia are related to human factors, primarily excessive water use in irrigated areas (Gupta et al. 2009 <sup>[[#fn:r1522|1522]]</sup> ), deep ploughing and lack of maintenance of vegetative cover in rainfed areas (Suleimenov et al. 2014 <sup>[[#fn:r1523|1523]]</sup> ), and overgrazing in rangelands (Mirzabaev et al. 2016 <sup>[[#fn:r1524|1524]]</sup> ). Lack of good maintenance of watering infrastructure for migratory livestock grazing, and fragmentation of livestock herds led to overgrazing near villages, increasing the soil erosion by wind (Alimaev et al. 2008 <sup>[[#fn:r1526|1526]]</sup> ). Overgrazing in the rangeland areas of the region (e.g., particularly in Kyzylkum) contributes to dust storms, coming primarily from the Ustyurt Plateau, desertified areas of Amudarya and Syrdarya rivers’ deltas, the dried seabed of the Aral Sea (now called Aralkum), and the Caspian Sea (Issanova and Abuduwaili 2017 <sup>[[#fn:r1527|1527]]</sup> ; Xi and Sokolik 2015). Xi and Sokolik (2015) estimated that total dust emissions in Central Asia were 255.6 Mt in 2001, representing 10–17% of the global total.

Central Asia is one of the regions highly exposed to climate change, with warming levels projected to be higher than the global mean (Hoegh-Guldberg et al. 2018 <sup>[[#fn:r1528|1528]]</sup> ), leading to more heat extremes (Reyer et al. 2017 <sup>[[#fn:r1529|1529]]</sup> ). There is no clear trend in precipitation extremes, with some potential for moderate rise in occurrence of droughts. The diminution of glaciers is projected to continue in the Pamir and Tian Shan mountain ranges, a major source of surface waters along with seasonal snowmelt. Glacier melting will increase the hazards from moraine-dammed glacial lakes and spring floods (Reyer et al. 2017 <sup>[[#fn:r1530|1530]]</sup> ). Increased intensity of spring floods creates favourable conditions for higher soil erosion by water, especially in the sloping areas in Kyrgyzstan and Tajikistan. The continuation of some of the current unsustainable cropland and rangeland management practices may lead to elevated rates of soil erosion, particularly in those parts of the region where climate change projections point to increases in floods (Kyrgyzstan, Tajikistan) or increases in droughts (Turkmenistan, Uzbekistan) (Hijioka et al. 2014 <sup>[[#fn:r1531|1531]]</sup> ). Increasing water use to compensate for higher evapotranspiration due to rising temperatures and heat waves could increase soil erosion by water in the irrigated zones, especially in sloping areas and crop fields with uneven land levelling (Bekchanov et al. 2010 <sup>[[#fn:r1532|1532]]</sup> ). The desiccation of the Aral Sea resulted in a hotter and drier regional microclimate, adding to the growing wind erosion in adjacent deltaic areas and deserts (Kust 1999 <sup>[[#fn:r1533|1533]]</sup> ).

There are numerous sustainable land and water management practices available in the region for reducing soil erosion (Abdullaev et al. 2007 <sup>[[#fn:r1534|1534]]</sup> ; Gupta et al. 2009 <sup>[[#fn:r1535|1535]]</sup> ; Kust et al. 2014 <sup>[[#fn:r1536|1536]]</sup> ; Nurbekov et al. 2016 <sup>[[#fn:r1537|1537]]</sup> ). These include: improved land levelling and more efficient irrigation methods such as drip, sprinkler and alternate furrow irrigation (Gupta et al. 2009 <sup>[[#fn:r1538|1538]]</sup> ); conservation agriculture practices, including no-till methods and maintenance of crop residues as mulch in the rainfed and irrigated areas (Kienzler et al. 2012 <sup>[[#fn:r1539|1539]]</sup> ; Pulatov et al. 2012 <sup>[[#fn:r1540|1540]]</sup> ); rotational grazing; institutional arrangements for pooling livestock for long-distance mobile grazing; reconstruction of watering infrastructure along the livestock migratory routes (Han et al. 2016; Mirzabaev et al. 2016 <sup>[[#fn:r1541|1541]]</sup> ); afforesting degraded marginal lands (Djanibekov and Khamzina 2016 <sup>[[#fn:r1543|1543]]</sup> ; Khamzina et al. 2009 <sup>[[#fn:r1545|1545]]</sup> ; Khamzina et al. 2016 <sup>[[#fn:r1546|1546]]</sup> ); integrated water resource management (Dukhovny et al. 2013 <sup>[[#fn:r1547|1547]]</sup> ; Kazbekov et al. 2009 <sup>[[#fn:r1548|1548]]</sup> ); and planting salt – and drought-tolerant halophytic plants as windbreaks in sandy rangelands (Akinshina et al. 2016 <sup>[[#fn:r1549|1549]]</sup> ; Qadir et al. 2009 <sup>[[#fn:r1550|1550]]</sup> ; Toderich et al. 2009 <sup>[[#fn:r1551|1551]]</sup> ; Toderich et al. 2008 <sup>[[#fn:r1552|1552]]</sup> ), and potentially the dried seabed of the former Aral Sea (Breckle 2013 <sup>[[#fn:r1553|1553]]</sup> ). The adoption of enabling policies, such as those discussed in Section 3.6.3, can facilitate the adoption of these sustainable land and water management practices in Central Asia ( ''high confidence'' ) (Aw-Hassan et al. 2016 <sup>[[#fn:r1554|1554]]</sup> ; Bekchanov et al. 2016 <sup>[[#fn:r1555|1555]]</sup> ; Bobojonov et al. 2013 <sup>[[#fn:r1556|1556]]</sup> ; Djanibekov et al. 2016 <sup>[[#fn:r1557|1557]]</sup> ; Hamidov et al. 2016 <sup>[[#fn:r1559|1559]]</sup> ; Mirzabaev et al. 2016 <sup>[[#fn:r1560|1560]]</sup> ).

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=== 3.7.2 Green walls and green dams ===

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This case study evaluates the experiences of measures and actions implemented to combat soil erosion, decrease dust storms, and to adapt to and mitigate climate change under the Green Wall and Green Dam programmes in East Asia (e.g., China) and Africa (e.g., Algeria, Sahara and the Sahel region). These measures have also been implemented in other countries, such as Mongolia (Do and Kang 2014; Lin et al. 2009), Turkey (Yurtoglu 2015 <sup>[[#fn:r1562|1562]]</sup> ; Çalişkan and Boydak 2017 <sup>[[#fn:r1563|1563]]</sup> ) and Iran (Amiraslani and Dragovich 2011 <sup>[[#fn:r1564|1564]]</sup> ), and are increasingly considered as part of many national and international initiatives to combat desertification (Goffner et al. 2019 <sup>[[#fn:r1565|1565]]</sup> ) (Cross-Chapter Box 2 in Chapter 1). Afforestation and reforestation programmes can contribute to reducing sand storms and increasing carbon sinks in dryland regions ( ''high confidence'' ). On the other hand, green wall and green dam programmes also decrease the albedo and hence increase the surface absorption of radiation, increasing the surface temperature. The net effect will largely depend on the balance between these and will vary from place to place depending on many factors.

<div id="section-3-7-2-1-the-experiences-of-combating-desertification-in-china"></div>

<span id="the-experiences-of-combating-desertification-in-china"></span>
==== 3.7.2.1 The experiences of combating desertification in China ====

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Arid and semi-arid areas of China, including north-eastern, northern and north-western regions, cover an area of more than 509 Mha, with annual rainfall of below 450 mm. Over the past several centuries, more than 60% of the areas in arid and semi-arid regions were used as pastoral and agricultural lands. The coupled impacts of past climate change and human activity have caused desertification and dust storms to become a serious problem in the region (Xu et al. 2010). In 1958, the Chinese government recognised that desertification and dust storms jeopardised the livelihoods of nearly 200 million people, and afforestation programmes for combating desertification have been initiated since 1978. China is committed to go beyond the Land Degradation Neutrality objective, as indicated by the following programmes that have been implemented. The Chinese Government began the Three North’s Forest Shelterbelt programme in Northeast China, North China, and Northwest China, with the goal to combat desertification and to control dust storms by improving forest cover in arid and semi-arid regions. The project is implemented in three stages (1978–2000, 2001–2020 and 2021–2050). In addition, the Chinese government launched the Beijing and Tianjin Sandstorm Source Treatment Project (2001–2010), Returning Farmlands to Forest Project (2003–present), and the Returning Grazing Land to Grassland Project (2003–present) to combat desertification, and for adaptation and mitigation of climate change (State Forestry Administration of China 2015 <sup>[[#fn:r1566|1566]]</sup> ; Wang 2014 <sup>[[#fn:r1567|1567]]</sup> ; Wang et al. 2013 <sup>[[#fn:r1568|1568]]</sup> ).

The results of the fifth monitoring period (2010–2014) showed: (i) compared with 2009, the area of degraded land decreased by 12,120 km <sup>2</sup> over a five-year period; (ii) in 2014, the average coverage of vegetation in the sand area was 18.33%, an increase of 0.7% compared with 17.63% in 2009, and the carbon sequestration increased by 8.5%; (iii) compared with 2009, the amount of wind erosion decreased by 33%, the average annual occurrence of sandstorms decreased by 20.3% in 2014; (iv) as of 2014, 203,700 km <sup>2</sup> of degraded land were effectively managed, accounting for 38.4% of the 530,000 km <sup>2</sup> of manageable desertified land; (v) the restoration of degraded land has created an annual output of 53.63 Mt of fresh and dried fruits, accounting for 33.9% of the total national annual output of fresh and dried fruits (State Forestry Administration of China 2015 <sup>[[#fn:r1570|1570]]</sup> ). This has become an important pillar for economic development and a high priority for peasants as a method to eradicate poverty (State Forestry Administration of China 2015 <sup>[[#fn:r1571|1571]]</sup> ).

Stable investment mechanisms for combating desertification have been established along with tax relief policies and financial support policies for guiding the country in its fight against desertification. The investments in scientific and technological innovation for combating desertification have been improved, the technologies for vegetation restoration under drought conditions have been developed, the popularisation and application of new technologies has been accelerated, and the training of technicians to assist farmers and herdsmen has been strengthened. To improve the monitoring capability and technical level of desertification studies, the monitoring network system has been strengthened, and the popularisation and application of modern technologies have been intensified (e.g., information technology and remote sensing) (Wu et al. 2015). Special laws on combating desertification have been decreed by the government. The provincial government’s responsibilities for desertification prevention and controlling objectives and laws have been strictly implemented.

Many studies showed that these projects generally played an active role in combating desertification and fighting against dust storms in China over the past several decades ( ''high confidence'' ) (Cao et al. 2018; State Forestry Administration of China 2015; Wang et al. 2013 <sup>[[#fn:r1573|1573]]</sup> ; Wang et al. 2014 <sup>[[#fn:r1574|1574]]</sup> ; Yang et al. 2013 <sup>[[#fn:r1576|1576]]</sup> ). At the beginning of the projects, some problems appeared in some places due to lack of enough knowledge and experience ( ''low confidence'' ) (Jiang 2016 <sup>[[#fn:r1578|1578]]</sup> ; Wang et al. 2010 <sup>[[#fn:r1579|1579]]</sup> ). For example, some tree species selected were not well suited to local soil and climatic conditions (Zhu et al. 2007), and there was inadequate consideration of the limitation of the amount of available water on the carrying capacity of trees in some arid regions (Dai 2011; Feng et al. 2016 <sup>[[#fn:r1580|1580]]</sup> ) (Section 3.6.4). In addition, at the beginning of the projects, there was an inadequate consideration of the effects of climate change on combating desertification (Feng et al. 2015 <sup>[[#fn:r1581|1581]]</sup> ; Tan and Li 2015). Indeed, climate change and human activities over past years have influenced the desertification and dust storm control effects in China (Feng et al. 2015 <sup>[[#fn:r1582|1582]]</sup> ; Wang et al. 2009 <sup>[[#fn:r1583|1583]]</sup> ; Tan and Li 2015), and future climate change will bring new challenges for combating desertification in China (Wang et al. 2017 <sup>[[#fn:r1584|1584]]</sup> ; Yin et al. 2015; Xu et al. 2019). In particular, the desertification risk in China will be enhanced at 2°C compared to 1.5°C global temperature rise (Ma et al. 2018). Adapting desertification control to climate change involves: improving the adaptation capacity to climate change for afforestation and grassland management by executing SLM practices; optimising the agricultural and animal husbandry structure; and using big data to meet the water resources regulation (Zhang and Huisingh 2018 <sup>[[#fn:r1588|1588]]</sup> ). In particular, improving scientific and technological supports in desertification control is crucial for adaptation to climate change and combating desertification, including protecting vegetation in desertification-prone lands by planting indigenous plant species, facilitating natural restoration of vegetation to conserve biodiversity, employing artificial rain or snow, water-saving irrigation and water storage technologies (Jin et al. 2014; Yang et al. 2013 <sup>[[#fn:r1589|1589]]</sup> ).

<div id="section-3-7-2-2-the-green-dam-in-algeria"></div>

<span id="the-green-dam-in-algeria"></span>
==== 3.7.2.2 The Green Dam in Algeria ====

<div id="section-3-7-2-2-the-green-dam-in-algeria-block-1"></div>

After independence in 1962, the Algerian government initiated measures to replant forests destroyed by the war, and the steppes affected by desertification, among its top priorities (Belaaz 2003 <sup>[[#fn:r1591|1591]]</sup> ).

In 1972, the government invested in the Green Dam ( ''Barrage vert'' ) project. This was the first significant experiment to combat desertification, influence the local climate and decrease the aridity by restoring a barrier of trees. The Green Dam extends across arid and semi-arid zones between the isohyets 300 mm and 200 mm. It is a 3 Mha band of plantation running from east to west (Figure 3.12). It is over 1200 km long (from the Algerian–Moroccan border to the Algerian–Tunisian border) and has an average width of about 20 km. The soils in the area are shallow, low in organic matter and susceptible to erosion. The main objectives of the project were to conserve natural resources, improve the living conditions of local residents and avoid their exodus to urban areas. During the first four decades (1970–2000) the success rate was low (42%) due to lack of participation by the local population and the choice of species (Bensaid 1995 <sup>[[#fn:r1592|1592]]</sup> ).

The Green Dam did not have the desired effects. Despite tree-planting efforts, desertification intensified on the steppes, especially in south-western Algeria, due to the prolonged drought during the 1980s. Rainfall declined in the range from 18% to 27%, and the dry season has increased by two months in the last century (Belala et al. 2018 <sup>[[#fn:r1593|1593]]</sup> ). Livestock numbers in the Green Dam regions, mainly sheep, grew exponentially, leading to severe overgrazing, causing trampling and soil compaction, which greatly increased the risk of erosion. Wind erosion, very prevalent in the region, is due to climatic conditions and the strong anthropogenic action that reduced the vegetation cover. The action of the wind carries fine particles such as sands and clays and leaves on the soil surface a lag-gravel pavement, which is unproductive. Water erosion is largely due to torrential rains in the form of severe thunderstorms that disintegrate the bare soil surface from raindrop impact (Achite et al. 2016 <sup>[[#fn:r1594|1594]]</sup> ). The detached soil and nutrients are transported offsite via runoff, resulting in loss of fertility and water holding capacity. The risk of and severity of water erosion is a function of human land-use activities that increase soil loss through removal of vegetative cover. The National Soil Sensitivity to Erosion Map (Salamani et al. 2012 <sup>[[#fn:r1595|1595]]</sup> ) shows that more than 3 Mha of land in the steppe provinces are currently experiencing intense wind activity (Houyou et al. 2016 <sup>[[#fn:r1596|1596]]</sup> ) and that these areas are at particular risk of soil erosion. Mostephaoui et al. (2013), estimates that each year there is a loss of 7 t ha <sup>–1</sup> of soils due to erosion. Nearly 0.6 Mha of land in the steppe zone are fully degraded without the possibility of biological recovery.

To combat the effects of erosion and desertification, the government has planned to relaunch the rehabilitation of the Green Dam by incorporating new concepts related to sustainable development, and adaptation to climate change. The experience of previous years has led to integrated rangeland management, improved tree and fodder shrub plantations and the development of water conservation techniques. Reforestation is carried out using several species, including fruit trees, to increase and diversify the sources of income for the population.

The evaluation of the Green Dam from 1972 to 2015 (Merdas et al. 2015 <sup>[[#fn:r1597|1597]]</sup> ) shows that 0.3 Mha of forest plantation have been planted, which represents 10% of the project area. Estimates of the success rate of reforestation vary considerably between 30% and 75%, depending on the region. Through demonstration, the Green Dam has inspired several African nations to work together to build a Great Green Wall to combat land degradation, mitigate climate change effects, loss of biodiversity and poverty in a region that stretches from Senegal to Djibouti (Sahara and Sahel Observatory (OSS) 2016) (Section 3.7.2.3).

<div id="section-3-7-2-2-the-green-dam-in-algeria-block-2"></div>

<span id="figure-3.12b"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.12b'''

<span id="location-of-the-green-dam-in-algeria-saifi-et-al.-2015.-note-the-green-coloured-band-represents-the-location-of-the-green-dam."></span>
<!-- IMG CAPTION -->
'''Location of the Green Dam in Algeria (Saifi et al. 2015). Note: The green coloured band represents the location of the Green Dam.'''
<!-- IMG FILE -->
[[File:852f131c5ef11fb337e11742bc73ccae Figure-3.12b-1024x577.jpg]]

Location of the Green Dam in Algeria (Saifi et al. 2015 <sup>[[#fn:r1806|1806]]</sup> ). Note: The green coloured band represents the location of the Green Dam.

<!-- END IMG -->
<div id="section-3-7-2-3-the-great-green-wall-of-the-sahara-and-the-sahel-initiative"></div>

<span id="the-great-green-wall-of-the-sahara-and-the-sahel-initiative"></span>
==== 3.7.2.3 The Great Green Wall of the Sahara and the Sahel Initiative ====

<div id="section-3-7-2-3-the-great-green-wall-of-the-sahara-and-the-sahel-initiative-block-1"></div>

The Great Green Wall is an initiative of the Heads of State and Government of the Sahelo-Saharan countries to mitigate and adapt to climate change, and to improve the food security of the Sahel and Saharan peoples (Sacande 2018 <sup>[[#fn:r1598|1598]]</sup> ; Mbow 2017 <sup>[[#fn:r1599|1599]]</sup> ). Launched in 2007, this regional project aims to restore Africa’s degraded arid landscapes, reduce the loss of biodiversity and support local communities to sustainable use of forests and rangelands. The Great Green Wall focuses on establishing plantations and neighbouring projects, covering a distance of 7775 km from Senegal on the Atlantic coast to Eritrea on the Red Sea coast, with a width of 15 km (Figure 3.13). The wall passes through Djibouti, Eritrea, Ethiopia, Sudan, Chad, Niger, Nigeria, Mali, Burkina Faso, Mauritania and Senegal.

The choice of woody and herbaceous species that will be used to restore degraded ecosystems is based on biophysical and socio-economic criteria, including socio-economic value (food, pastoral, commercial, energetic, medicinal, cultural); ecological importance (carbon sequestration, soil cover, water infiltration); and resilience to climate change and variability. The Pan-African Agency of the Great Green Wall (PAGGW) was created in 2010 under the auspices of the African Union and CEN-SAD to manage the project. The initiative is implemented at the level of each country by a national structure. A monitoring and evaluation system has been defined, allowing nations to measure outcomes and to propose the necessary adjustments.

In the past, reforestation programmes in the arid regions of the Sahel and North Africa that have been undertaken to stop desertification were poorly studied and cost a lot of money without significant success (Benjaminsen and Hiernaux 2019 <sup>[[#fn:r1600|1600]]</sup> ). Today, countries have changed their strategies and opted for rural development projects that can be more easily funded. Examples of scalable practices for land restoration include managing water bodies for livestock and crop production, and promoting fodder trees to reduce runoff (Mbow 2017 <sup>[[#fn:r1601|1601]]</sup> ).

The implementation of the initiative has already started in several countries. For example, the FAO’s Action Against Desertification project was restoring 18,000 hectares of land in 2018 through planting native tree species in Burkina Faso, Ethiopia, The Gambia, Niger, Nigeria and Senegal (Sacande 2018 <sup>[[#fn:r1602|1602]]</sup> ). Berrahmouni et al. (2016) <sup>[[#fn:r1807|1807]]</sup> estimated that 166 Mha can be restored in the Sahel, requiring the restoration of 10 Mha per year to achieve Land Degradation Neutrality targets by 2030. Despite these early implementation actions on the ground, the achievement of the planned targets is questionable, and will be challenging without significant additional funding.

<div id="section-3-7-2-3-the-great-green-wall-of-the-sahara-and-the-sahel-initiative-block-2"></div>

<span id="figure-3.13"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.13'''

<span id="the-great-green-wall-of-the-sahara-and-the-sahel.-source-for-the-data-layer-this-dataset-is-an-extract-from-the-globcover-2009-land-cover-map-covering-africa-and-the-arabian-peninsula.-the-globcover-2009-land-cover-map-is-derived-by-an-automatic-and-regionally-tuned-classification-of-a-time-series-of-global-meris"></span>
<!-- IMG CAPTION -->
'''The Great Green Wall of the Sahara and the Sahel. Source for the data layer: This dataset is an extract from the GlobCover 2009 land cover map, covering Africa and the Arabian Peninsula. The GlobCover 2009 land cover map is derived by an automatic and regionally tuned classification of a time series of global MERIS […]'''
<!-- IMG FILE -->
[[File:7908b973d5538d2af6a02e2650686e0f Figure-3.13.jpg]]

The Great Green Wall of the Sahara and the Sahel. Source for the data layer: This dataset is an extract from the GlobCover 2009 land cover map, covering Africa and the Arabian Peninsula. The GlobCover 2009 land cover map is derived by an automatic and regionally tuned classification of a time series of global MERIS (MEdium Resolution Imaging Spectrometer) FR mosaics for the year 2009. The global land cover map counts 22 land cover classes defined with the United Nations (UN) Land Cover Classification System (LCCS)

<!-- END IMG -->
<span id="invasive-plant-species"></span>
=== 3.7.3 Invasive plant species ===

<div id="section-3-7-3-1-introduction"></div>

<span id="introduction-1"></span>
==== 3.7.3.1 Introduction ====

<div id="section-3-7-3-1-introduction-block-1"></div>

The spread of invasive plants can be exacerbated by climate change (Bradley et al. 2010 <sup>[[#fn:r1603|1603]]</sup> ; Davis et al. 2000 <sup>[[#fn:r1604|1604]]</sup> ). In general, it is expected that the distribution of invasive plant species with high tolerance to drought or high temperatures may increase under most climate change scenarios ( ''medium to high confidence'' ) (Bradley et al. 2010 <sup>[[#fn:r1605|1605]]</sup> ; Settele et al. 2014 <sup>[[#fn:r1606|1606]]</sup> ; Scasta et al. 2015 <sup>[[#fn:r1607|1607]]</sup> ). Invasive plants are considered a major risk to native biodiversity and can disturb the nutrient dynamics and water balance in affected ecosystems (Ehrenfeld 2003 <sup>[[#fn:r1608|1608]]</sup> ). Compared to more humid regions, the number of species that succeed in invading dryland areas is low (Bradley et al. 2012 <sup>[[#fn:r1609|1609]]</sup> ), yet they have a considerable impact on biodiversity and ecosystem services (Le Maitre et al. 2015, 2011; Newton et al. 2011 <sup>[[#fn:r1610|1610]]</sup> ). Moreover, human activities in dryland areas are responsible for creating new invasion opportunities (Safriel et al. 2005 <sup>[[#fn:r1611|1611]]</sup> ).

Current drivers of species introductions include expanding global trade and travel, land degradation and changes in climate (Chytrý et al. 2012 <sup>[[#fn:r1612|1612]]</sup> ; Richardson et al. 2011 <sup>[[#fn:r1613|1613]]</sup> ; Seebens et al. 2018 <sup>[[#fn:r1614|1614]]</sup> ). For example, Davis et al. (2000) suggests that high rainfall variability promotes the success of alien plant species – as reported for semi-arid grasslands and Mediterranean-type ecosystems (Cassidy et al. 2004 <sup>[[#fn:r1615|1615]]</sup> ; Reynolds et al. 2004 <sup>[[#fn:r1616|1616]]</sup> ; Sala et al. 2006 <sup>[[#fn:r1617|1617]]</sup> ). Furthermore, Panda et al. (2018) demonstrated that many invasive species could withstand elevated temperature and moisture scarcity caused by climate change. Dukes et al. (2011) observed that the invasive plant yellow-star thistle ( ''Centaurea solstitialis'' ) grew six time larger under the elevated atmospheric CO <sub>2</sub> expected in future climate change scenarios.

Climate change is ''likely'' to aggravate the problem as existing species continue to spread unabated and other species develop invasive characteristics (Hellmann et al. 2008 <sup>[[#fn:r1619|1619]]</sup> ). Although the effects of climate change on invasive species distributions have been relatively well explored, the greater impact on ecosystems is less well understood (Bradley et al. 2010 <sup>[[#fn:r1620|1620]]</sup> ; Eldridge et al. 2011 <sup>[[#fn:r1621|1621]]</sup> ).

Due to the time lag between the initial release of invasive species and their impact, the consequence of invasions is not immediately detected and may only be noticed centuries after introduction (Rouget et al. 2016 <sup>[[#fn:r1622|1622]]</sup> ). Climate change and invading species may act in concert (Bellard et al. 2013 <sup>[[#fn:r1623|1623]]</sup> ; Hellmann et al. 2008 <sup>[[#fn:r1625|1625]]</sup> ; Seebens et al. 2015 <sup>[[#fn:r1626|1626]]</sup> ). For example, invasion often changes the size and structure of fuel loads, which can lead to an increase in the frequency and intensity of fire (Evans et al. 2015). In areas where the climate is becoming warmer, an increase in the likelihood of suitable weather conditions for fire may promote invasive species, which in turn may lead to further desertification. Conversely, fire may promote plant invasions via several mechanisms (by reducing cover of competing vegetation, destroying native vegetation and clearing a path for invasive plants or creating favourable soil conditions) (Brooks et al. 2004 <sup>[[#fn:r1627|1627]]</sup> ; Grace et al. 2001 <sup>[[#fn:r1628|1628]]</sup> ; Keeley and Brennan 2012 <sup>[[#fn:r1629|1629]]</sup> ).

<div id="section-3-7-3-1-introduction-block-2"></div>

<span id="figure-3.14"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.14'''

<span id="difference-between-the-number-of-invasive-alien-species-n99-from-bellard-et-al.-2013-predicted-to-occur-by-2050-under-a1b-scenario-and-current-period-2000-within-the-dryland-areas"></span>
<!-- IMG CAPTION -->
'''Difference between the number of invasive alien species (n=99, from Bellard et al. (2013)) predicted to occur by 2050 (under A1B scenario) and current period ‘2000’ within the dryland areas'''
<!-- IMG FILE -->
[[File:f7e78c29625c2d159d611f7fff53955a Figure-3.14.png]]

Difference between the number of invasive alien species (n=99, from Bellard et al. (2013) <sup>[[#fn:r1808|1808]]</sup> ) predicted to occur by 2050 (under A1B scenario) and current period ‘2000’ within the dryland areas

<!-- END IMG -->
<div id="section-3-7-3-1-introduction-block-3"></div>

At a regional scale, Bellard et al. (2013) <sup>[[#fn:r1809|1809]]</sup> predicted increasing risk in Africa and Asia, with declining risk in Australia (Figure 3.14). This projection does not represent an exhaustive list of invasive alien species occurring in drylands.

A set of four case studies in Ethiopia, Mexico, the USA and Pakistan is presented below to describe the nuanced nature of invading plant species, their impact on drylands and their relationship with climate change.

<div id="section-3-7-3-2-ethiopia"></div>

<span id="ethiopia"></span>
==== 3.7.3.2 Ethiopia ====

<div id="section-3-7-3-2-ethiopia-block-1"></div>

The two invasive plants that inflict the heaviest damage to ecosystems, especially biodiversity, are the annual herbaceous weed, ''Parthenium hysterophorus'' ( ''Asteraceae'' ) also known as Congress weed; and the tree species, ''Prosopis juliflora (Fabaceae'' ) also called Mesquite, both originating from the southwestern United States to Central/South America (Adkins and Shabbir 2014 <sup>[[#fn:r1630|1630]]</sup> ). ''Prosopis'' was introduced in the 1970s and has since spread rapidly. ''Prosopis'' , classified as the highest priority invader in Ethiopia, is threatening livestock production and challenging the sustainability of the pastoral systems. ''Parthenium'' is believed to have been introduced along with relief aid during the debilitating droughts of the early 1980s, and a recent study reported that it has spread into 32 out of 34 districts in Tigray, the northernmost region of Ethiopia (Teka 2016 <sup>[[#fn:r1631|1631]]</sup> ). A study by Etana et al. (2011) indicated that Parthenium caused a 69% decline in the density of herbaceous species in Awash National Park within a few years of introduction. In the presence of Parthenium, the growth and development of crops is suppressed due to its allelopathic properties. McConnachie et al. (2011) estimated a 28% crop loss across the country, including a 40–90% reduction in sorghum yield in eastern Ethiopia alone (Tamado et al. 2002 <sup>[[#fn:r1632|1632]]</sup> ). The weed is a substantial agricultural and natural resource problem and constitutes a significant health hazard (Fasil 2011). Parthenium causes acute allergic respiratory problems, skin dermatitis, and reportedly mutagenicity both in humans and livestock (Mekonnen 2017; Patel 2011 <sup>[[#fn:r1633|1633]]</sup> ). The eastern belt of Africa – including Ethiopia – presents a very suitable habitat, and the weed is expected to spread further in the region in the future (Mainali et al. 2015 <sup>[[#fn:r1635|1635]]</sup> ).

There is neither a comprehensive intervention plan nor a clear institutional mandate to deal with invasive weeds, however, there are fragmented efforts involving local communities even though they are clearly inadequate. The lessons learned, related to actions that have contributed to the current scenario, are several. First, lack of coordination and awareness – mesquite was introduced by development agencies as a drought-tolerant shade tree with little consideration of its invasive nature. If research and development institutions had been aware, a containment strategy could have been implemented early on. The second major lesson is the cost of inaction. When research and development organisations did sound the alarm, the warnings went largely unheeded, resulting in the spread and buildup of two of the worst invasive plant species in the world (Fasil 2011 <sup>[[#fn:r1636|1636]]</sup> ).

<div id="section-3-7-3-3-mexico"></div>

<span id="mexico"></span>
==== 3.7.3.3 Mexico ====

<div id="section-3-7-3-3-mexico-block-1"></div>

Buffelgrass ( ''Cenchrus ciliaris'' L.), a native species from southern Asia and East Africa, was introduced into Texas and northern Mexico in the 1930s and 1940s, as it is highly productive in drought conditions (Cox et al. 1988; Rao et al. 1996). In the Sonoran desert of Mexico, the distribution of buffelgrass has increased exponentially, covering 1 Mha in Sonora State (Castellanos-Villegas et al. 2002 <sup>[[#fn:r1637|1637]]</sup> ). Furthermore, its potential distribution extended to 53% of Sonora State and 12% of semi-arid and arid ecosystems in Mexico (Arriaga et al. 2004 <sup>[[#fn:r1638|1638]]</sup> ). Buffelgrass has also been reported as an aggressive invader in Australia and the USA, resulting in altered fire cycles that enhance further spread of this plant and disrupt ecosystem processes (Marshall et al. 2012 <sup>[[#fn:r1639|1639]]</sup> ; Miller et al. 2010 <sup>[[#fn:r1641|1641]]</sup> ; Schlesinger et al. 2013 <sup>[[#fn:r1642|1642]]</sup> ).

Castellanos et al. (2016) reported that soil moisture was lower in the buffelgrass savannah cleared 35 years ago than in the native semi-arid shrubland, mainly during the summer. The ecohydrological changes induced by buffelgrass can therefore displace native plant species over the long term. Invasion by buffelgrass can also affect landscape productivity, as it is not as productive as native vegetation (Franklin and Molina-Freaner 2010 <sup>[[#fn:r1643|1643]]</sup> ). Incorporation of buffelgrass is considered a good management practice by producers and the government. For this reason, no remedial actions are undertaken.

<div id="section-3-7-3-4-united-states-of-america"></div>

<span id="united-states-of-america"></span>
==== 3.7.3.4 United States of America ====

<div id="section-3-7-3-4-united-states-of-america-block-1"></div>

Sagebrush ecosystems have declined from 25 Mha to 13 Mha since the late 1800s (Miller et al. 2011 <sup>[[#fn:r1644|1644]]</sup> ). A major cause is the introduction of non-native cheatgrass ( ''Bromus tectorum'' ), which is the most prolific invasive plant in the USA. Cheatgrass infests more than 10 Mha in the Great Basin and is expanding every year (Balch et al. 2013 <sup>[[#fn:r1645|1645]]</sup> ). It provides a fine-textured fuel that increases the intensity, frequency and spatial extent of fire (Balch et al. 2013). Historically, wildfire frequency was 60 to 110 years in Wyoming big sagebrush communities and has increased to five years following the introduction of cheatgrass (Balch et al. 2013 <sup>[[#fn:r1646|1646]]</sup> ; Pilliod et al. 2017 <sup>[[#fn:r1648|1648]]</sup> ).

The conversion of the sagebrush steppe biome to annual grassland with higher fire frequencies has severely impacted livestock producers, as grazing is not possible for a minimum of two years after fire. Furthermore, cheatgrass and wildfires reduce critical habitat for wildlife and negatively impact species richness and abundance – for example, the greater sage-grouse ( ''Centocercus urophasianus'' ) and pygmy rabbit ( ''Brachylagus idahoensis'' ) which are on the verge of being listed for federal protection (Crawford et al. 2004 <sup>[[#fn:r1649|1649]]</sup> ; Larrucea and Brussard 2008 <sup>[[#fn:r1650|1650]]</sup> ; Lockyer et al. 2015 <sup>[[#fn:r1651|1651]]</sup> ).

Attempts to reduce cheatgrass impacts through reseeding of both native and adapted introduced species have occurred for more than 60 years (Hull and Stewart 1949 <sup>[[#fn:r1652|1652]]</sup> ) with little success. Following fire, cheatgrass becomes dominant and recovery of native shrubs and grasses is improbable, particularly in relatively low-elevation sites with minimal annual precipitation (less than 200 mm yr <sup>–1</sup> ) (Davies et al. 2012 <sup>[[#fn:r1653|1653]]</sup> ; Taylor et al. 2014 <sup>[[#fn:r1654|1654]]</sup> ). Current rehabilitation efforts emphasise the use of native and non-native perennial grasses, forbs and shrubs (Bureau of Land Management 2005 <sup>[[#fn:r1655|1655]]</sup> ). Recent literature suggests that these treatments are not consistently effective at displacing cheatgrass populations or re-establishing sage-grouse habitat, with success varying with elevation and precipitation (Arkle et al. 2014 <sup>[[#fn:r1656|1656]]</sup> ; Knutson et al. 2014 <sup>[[#fn:r1657|1657]]</sup> ). Proper post-fire grazing rest, season-of-use, stocking rates, and subsequent management are essential to restore resilient sagebrush ecosystems before they cross a threshold and become an annual grassland (Chambers et al. 2014 <sup>[[#fn:r1658|1658]]</sup> ; Miller et al. 2011 <sup>[[#fn:r1659|1659]]</sup> ; Pellant et al. 2004 <sup>[[#fn:r1660|1660]]</sup> ). Biological soil crust protection may be an effective measure to reduce cheatgrass germination, as biocrust disturbance has been shown to be a key factor promoting germination of non-native grasses (Hernandez and Sandquist 2011). Projections of increasing temperature (Abatzoglou and Kolden 2011 <sup>[[#fn:r1662|1662]]</sup> ), and observed reductions in and earlier melting of snowpack in the Great Basin region (Harpold and Brooks 2018 <sup>[[#fn:r1663|1663]]</sup> ; Mote et al. 2005 <sup>[[#fn:r1664|1664]]</sup> ) suggest that there is a need to understand current and past climatic variability as this will drive wildfire variability and invasions of annual grasses.

<div id="section-3-7-3-5-pakistan"></div>

<span id="pakistan"></span>
==== 3.7.3.5 Pakistan ====

<div id="section-3-7-3-5-pakistan-block-1"></div>

The alien plants invading local vegetation in Pakistan include ''Brossentia papyrifera'' (found in Islamabad Capital territory), ''Parthenium hysterophorus'' (found in Punjab and Khyber Pakhtunkhwa provinces), ''Prosopis juliflora'' (found all over Pakistan), ''Eucalyptus camaldulensis'' (found in Punjab and Sindh provinces), ''Salvinia'' (aquatic plant widely distributed in water bodies in Sindh), ''Cannabis sativa'' (found in Islamabad Capital Territory), ''Lantana camara'' and ''Xanthium strumarium'' (found in upper Punjab and Khyber Pakhtunkhwa provinces) (Khan et al. 2010 <sup>[[#fn:r1665|1665]]</sup> ; Qureshi et al. 2014 <sup>[[#fn:r1666|1666]]</sup> ). Most of these plants were introduced by the Forest Department decades ago for filling the gap between demand and supply of timber, fuelwood and fodder. These non-native plants have some uses but their disadvantages outweigh their benefits (Marwat et al. 2010 <sup>[[#fn:r1667|1667]]</sup> ; Rashid et al. 2014 <sup>[[#fn:r1668|1668]]</sup> ).

Besides being a source of biological pollution and a threat to biodiversity and habitat loss, the alien plants reduce the land value and cause huge losses to agricultural communities (Rashid et al. 2014 <sup>[[#fn:r1810|1810]]</sup> ). ''Brossentia papyrifera'' , commonly known as Paper Mulberry, is the root cause of inhalant pollen allergy for the residents of lush green Islamabad during spring. From February to April, the pollen allergy is at its peak, with symptoms of severe persistent coughing, difficulty in breathing, and wheezing. The pollen count, although variable at different times and days, can be as high as 55,000 m <sup>-3</sup> .

Early symptoms of the allergy include sneezing, itching in the eyes and skin, and blocked nose. With changing climate, the onset of disease is getting earlier, and pollen count is estimated to cross 55,000 m <sup>–3</sup> (Rashid et al. 2014 <sup>[[#fn:r1670|1670]]</sup> ). About 45% of allergic patients in the twin cities of Islamabad and Rawalpindi showed positive sensitivity to the pollens (Marwat et al. 2010 <sup>[[#fn:r1671|1671]]</sup> ). Millions of rupees have been spent by the Capital Development Authority on pruning and cutting of Paper Mulberry trees but because of its regeneration capacity growth is regained rapidly (Rashid et al. 2014 <sup>[[#fn:r1672|1672]]</sup> ). Among other invading plants, ''Prosopis juliflora'' has allelopathic properties, and ''Eucalyptus'' is known to transpire huge amounts of water and deplete the soil of its nutrient elements (Qureshi et al. 2014 <sup>[[#fn:r1673|1673]]</sup> ).

Although a Biodiversity Action Plan exists in Pakistan, it is not implemented in letter or spirit. The Quarantine Department focuses only on pests and pathogens but takes no notice of plant and animal species being imported. Also, there is no provision for checking the possible impacts of imported species on the environment (Rashid et al. 2014 <sup>[[#fn:r1674|1674]]</sup> ) or for carrying out bioassays of active allelopathic compounds of alien plants.

<span id="oases-in-hyper-arid-areas-in-the-arabian-peninsula-and-northern-africa"></span>
=== 3.7.4 Oases in hyper-arid areas in the Arabian Peninsula and northern Africa ===

<div id="section-3-7-4-oases-in-hyper-arid-areas-in-the-arabian-peninsula-and-northern-africa-block-1"></div>

Oases are isolated areas with reliable water supply from lakes and springs, located in hyper-arid and arid zones (Figure 3.15). Oasis agriculture has long been the only viable crop production system throughout the hot and arid regions of the Arabian Peninsula and North Africa. Oases in hyper-arid climates are usually subject to water shortage as evapotranspiration exceeds rainfall. This often causes salinisation of soils. While many oases have persisted for several thousand years, many others have been abandoned, often in response to changes in climate or hydrologic conditions (Jones et al. 2019 <sup>[[#fn:r1675|1675]]</sup> ), providing testimony to societies’ vulnerability to climatic shifts and raising concerns about similarly severe effects of anthropogenic climate change (Jones et al. 2019 <sup>[[#fn:r1676|1676]]</sup> ).

On the Arabian Peninsula and in North Africa, climate change is projected to have substantial and complex effects on oasis areas (Abatzoglou and Kolden 2011 <sup>[[#fn:r1677|1677]]</sup> ; Ashkenazy et al. 2012 <sup>[[#fn:r1678|1678]]</sup> ; Bachelet et al. 2016 <sup>[[#fn:r1679|1679]]</sup> ; Guan et al. 2018 <sup>[[#fn:r1680|1680]]</sup> ; Iknayan and Beissinger 2018 <sup>[[#fn:r1681|1681]]</sup> ; Ling et al. 2013 <sup>[[#fn:r1682|1682]]</sup> ). To illustrate, by the 2050s, the oases in southern Tunisia are expected to be affected by hydrological and thermal changes, with an average temperature increase of 2.7°C, a 29% decrease in precipitation and a 14% increase in evapotranspiration rate (Ministry of Agriculture and Water Resources of Tunisia and GIZ 2007 <sup>[[#fn:r1683|1683]]</sup> ). In Morocco, declining aquifer recharge is expected to impact the water supply of the Figuig oasis (Jilali 2014 <sup>[[#fn:r1684|1684]]</sup> ), as well as for the Draa Valley (Karmaoui et al. 2016 <sup>[[#fn:r1685|1685]]</sup> ). Saudi Arabia is expected to experience a 1.8°C–4.1°C increase in temperatures by 2050, which is forecast to raise agricultural water demand by 5–15% in order to maintain production levels equal to those of 2011 (Chowdhury and Al-Zahrani 2013 <sup>[[#fn:r1686|1686]]</sup> ). The increase of temperatures and variable pattern of rainfall over the central, north and south-western regions of Saudi Arabia may pose challenges for sustainable water resource management (Tarawneh and Chowdhury 2018 <sup>[[#fn:r1687|1687]]</sup> ). Moreover, future climate scenarios are expected to increase the frequency of floods and flash floods, such as in the coastal areas along the central parts of the Red Sea and the south-southwestern areas of Saudi Arabia (Almazroui et al. 2017 <sup>[[#fn:r1688|1688]]</sup> ).

While many oases are cultivated with very heat-tolerant crops such as date palms, even such crops eventually have declines in their productivity when temperatures exceed certain thresholds or hot conditions prevail for extended periods. Projections so far do not indicate severe losses in land suitability for date palm for the Arabian Peninsula (Aldababseh et al. 2018 <sup>[[#fn:r1689|1689]]</sup> ; Shabani et al. 2015 <sup>[[#fn:r1690|1690]]</sup> ). It is unclear, however, how reliable the climate response parameters in the underlying models are, and actual responses may differ substantially.

<div id="section-3-7-4-oases-in-hyper-arid-areas-in-the-arabian-peninsula-and-northern-africa-block-2"></div>

<span id="figure-3.15a"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.15a'''

<span id="oases-across-the-arabian-peninsula-and-north-africa-alphabetically-by-country.-a-masayrat-ar-ruwajah-oasis-ad-dakhiliyah-governorate-oman-photo-eike-lüdeling."></span>
<!-- IMG CAPTION -->
'''Oases across the Arabian Peninsula and North Africa (alphabetically by country). (a) Masayrat ar Ruwajah oasis, Ad Dakhiliyah Governorate, Oman (Photo: Eike Lüdeling).'''
<!-- IMG FILE -->
[[File:c5b723bb68f976aacd770713c3a40f92 Figure-3.15a-1024x591.jpg]]

Oases across the Arabian Peninsula and North Africa (alphabetically by country). (a) Masayrat ar Ruwajah oasis, Ad Dakhiliyah Governorate, Oman (Photo: Eike Lüdeling).

<!-- END IMG -->
<div id="section-3-7-4-oases-in-hyper-arid-areas-in-the-arabian-peninsula-and-northern-africa-block-3"></div>

<span id="figure-3.15b"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.15b'''

<span id="b-tasselmanet-oasis-ouarzazate-province-morocco-photo-abdellatif-khattabi."></span>
<!-- IMG CAPTION -->
'''(b) Tasselmanet oasis, Ouarzazate Province, Morocco (Photo: Abdellatif Khattabi).'''
<!-- IMG FILE -->
[[File:cb2462b9367442135d266b30fc0ff113 Figure-3.15b-1024x768.jpg]]

(b) Tasselmanet oasis, Ouarzazate Province, Morocco (Photo: Abdellatif Khattabi).

<!-- END IMG -->
<div id="section-3-7-4-oases-in-hyper-arid-areas-in-the-arabian-peninsula-and-northern-africa-block-4"></div>

<span id="figure-3.15c"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.15c'''

<span id="c-al-ahsa-oasis-al-ahsa-governarate-saudi-arabia-photo-shijan-kaakkara."></span>
<!-- IMG CAPTION -->
'''c) Al-Ahsa oasis, Al-Ahsa Governarate, Saudi Arabia (Photo: Shijan Kaakkara).'''
<!-- IMG FILE -->
[[File:4273257b16d71253d76d69ad3053be32 Figure-3.15c-1024x468.jpg]]

c) Al-Ahsa oasis, Al-Ahsa Governarate, Saudi Arabia (Photo: Shijan Kaakkara).

<!-- END IMG -->
<div id="section-3-7-4-oases-in-hyper-arid-areas-in-the-arabian-peninsula-and-northern-africa-block-5"></div>

<span id="figure-3.15d"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure-3.15d'''

<span id="zarat-oasis-governorate-of-gabes-tunisia-photo-hamda-aloui.-the-use-rights-for-a-b-and-d-were-granted-by-copyright-holders-c-is-licensed-under-the-creative-commons-attribution-2.0-generic-license."></span>
<!-- IMG CAPTION -->
'''Zarat oasis, Governorate of Gabes, Tunisia (Photo: Hamda Aloui). The use rights for (a), (b) and (d) were granted by copyright holders; (c) is licensed under the Creative Commons Attribution 2.0 Generic license.'''
<!-- IMG FILE -->
[[File:60be05098aa58f7a2c253bef8575e3b1 Figure-3.15d.jpg]]

Zarat oasis, Governorate of Gabes, Tunisia (Photo: Hamda Aloui). The use rights for (a), (b) and (d) were granted by copyright holders; (c) is licensed under the Creative Commons Attribution 2.0 Generic license.

<!-- END IMG -->
<div id="section-3-7-4-oases-in-hyper-arid-areas-in-the-arabian-peninsula-and-northern-africa-block-6"></div>

Date palms are routinely assumed to be able to endure very high temperatures, but recent transcriptomic and metabolomic evidence suggests that heat stress reactions already occur at 35°C (Safronov et al. 2017 <sup>[[#fn:r1691|1691]]</sup> ), which is not exceptionally warm for many oases in the region. Given current assumptions about the heat-tolerance of date palm, however, adverse effects are expected to be small (Aldababseh et al. 2018 <sup>[[#fn:r1692|1692]]</sup> ; Shabani et al. 2015 <sup>[[#fn:r1693|1693]]</sup> ). For some other perennial oasis crops, impacts of temperature increases are already apparent. Between 2004/2005 and 2012/2013, high-mountain oases of Al Jabal Al Akhdar in Oman lost almost all fruit and nut trees of temperate-zone origin, with the abundance of peaches, apricots, grapes, figs, pears, apples, and plums dropping by between 86% and 100% (Al-Kalbani et al. 2016 <sup>[[#fn:r1694|1694]]</sup> ). This implies that that the local climate may not remain suitable for species that depend on cool winters to break their dormancy period (Luedeling et al. 2009 <sup>[[#fn:r1695|1695]]</sup> ). A similar impact is very probable in Tunisia and Morocco, as well as in other oasis locations in the Arabian Peninsula and North Africa (Benmoussa et al. 2007 <sup>[[#fn:r1811|1811]]</sup> ). All these studies expect strong decreases in winter chill, raising concerns that many currently well-established species will no longer be viable in locations where they are grown today. The risk of detrimental chill shortfalls is expected to increase gradually, slowly diminishing the economic prospects to produce such species. Without adequate adaptation actions, the consequences of this development for many traditional oasis settlements and other plantations of similar species could be highly negative.

At the same time, population growth and agricultural expansion in many oasis settlements are leading to substantial increases in water demand for human consumption (Al-Kalbani et al. 2014 <sup>[[#fn:r1696|1696]]</sup> ). For example, a large unmet water demand has been projected for future scenarios in the valley of Seybouse in East Algeria (Aoun-Sebaiti et al. 2014 <sup>[[#fn:r1697|1697]]</sup> ), and similar conclusions were drawn for Wadi El Natrun in Egypt (Switzman et al. 2018 <sup>[[#fn:r1698|1698]]</sup> ). Modelling studies have indicated long-term decline in available water and increasing risk of water shortages – for example, for oases in Morocco (Johannsen et al. 2016 <sup>[[#fn:r1699|1699]]</sup> ; Karmaoui et al. 2016 <sup>[[#fn:r1700|1700]]</sup> ), the Dakhla oasis in Egypt’s Western Desert (Sefelnasr et al. 2014 <sup>[[#fn:r1701|1701]]</sup> ) and for the large Upper Mega Aquifer of the Arabian Peninsula (Siebert et al. 2016 <sup>[[#fn:r1702|1702]]</sup> ). Mainly due to the risk of water shortages, Souissi et al. (2018) classified almost half of all farmers in Tunisia as non-resilient to climate change, especially those relying on tree crops, which limit opportunities for short-term adaptation actions.

The maintenance of the oasis systems and the safeguarding of their population’s livelihoods are currently threatened by continuous water degradation, increasing soil salinisation, and soil contamination (Besser et al. 2017 <sup>[[#fn:r1703|1703]]</sup> ). Waterlogging and salinisation of soils due to rising saline groundwater tables coupled with inefficient drainage systems have become common to all continental oases in Tunisia, most of which are concentrated around saline depressions, known locally as chotts (Ben Hassine et al. 2013 <sup>[[#fn:r1704|1704]]</sup> ). Similar processes of salinisation are also occurring in the oasis areas of Egypt due to agricultural expansion, excessive use of water for irrigation and deficiency of the drainage systems (Abo-Ragab 2010 <sup>[[#fn:r1705|1705]]</sup> ; Masoud and Koike 2006 <sup>[[#fn:r1706|1706]]</sup> ). A prime example for this is Siwa oasis (Figure 3.16), a depression extending over 1050 km <sup>2</sup> in the north-western desert of Egypt in the north of the sand dune belt of the Great Sand Sea (Abo-Ragab and Zaghloul 2017 <sup>[[#fn:r1707|1707]]</sup> ). Siwa oasis has been recognised as a Globally Important Agricultural Heritage Site (GIAHS) by the FAO for being an ''in situ'' repository of plant genetic resources, especially of uniquely adapted varieties of date palm, olive and secondary crops that are highly esteemed for their quality and continue to play a significant role in rural livelihoods and diets (FAO 2016).

The population growth in Siwa is leading rapid agricultural expansion and land reclamation.The Siwan farmers are converting the surrounding desert into reclaimed land by applying their old inherited traditional practices. Yet, agricultural expansion in the oasis mainly depends on non-renewable groundwaters. Soil salinisation and vegetation loss have been accelerating since 2000 due to water mismanagement and improper drainage systems (Masoud and Koike 2006 <sup>[[#fn:r1708|1708]]</sup> ). Between 1990 and 2008, the cultivated area increased from 53 to 88 km , lakes from 60 to 76 km <sup>2</sup> , ''sabkhas'' (salt flats) from 335 to 470 km <sup>2</sup> , and the urban area from 6 to 10 km <sup>2</sup> (Abo-Ragab 2010 <sup>[[#fn:r1709|1709]]</sup> ). The problem of rising groundwater tables was exacerbated by climatic changes (Askri et al. 2010 <sup>[[#fn:r1710|1710]]</sup> ; Gad and Abdel-Baki 2002; Marlet et al. 2009 <sup>[[#fn:r1711|1711]]</sup> ).

Water supply is ''likely'' to become even scarcer for oasis agriculture under changing climate in the future than it is today, and viable solutions are difficult to find. While some authors stress the possibility to use desalinated water for irrigation (Aldababseh et al. 2018 <sup>[[#fn:r1712|1712]]</sup> ), the economics of such options, especially given the high evapotranspiration rates in the Arabian Peninsula and North Africa, are debatable. Many oases are located far from water sources that are suitable for desalination, adding further to feasibility constraints. Most authors therefore stress the need to limit water use (Sefelnasr et al. 2014 <sup>[[#fn:r1713|1713]]</sup> ), for example, by raising irrigation efficiency (Switzman et al. 2018 <sup>[[#fn:r1714|1714]]</sup> ), reducing agricultural areas (Johannsen et al. 2016 <sup>[[#fn:r1715|1715]]</sup> ) or imposing water use restrictions (Odhiambo 2017 <sup>[[#fn:r1716|1716]]</sup> ), and to carefully monitor desertification (King and Thomas 2014 <sup>[[#fn:r1717|1717]]</sup> ). Whether adoption of crops with low water demand, such as sorghum ( ''Sorghum bicolor'' (L.) Moench) or jojoba ( ''Simmondsia chinensis'' (Link) C. K. Schneid.) (Aldababseh et al. 2018 <sup>[[#fn:r1718|1718]]</sup> ), can be a viable option for some oases remains to be seen, but given their relatively low profit margins compared to currently grown oasis crops, there are reasons to doubt the economic feasibility of such proposals. While it is currently unclear to what extent oasis agriculture can be maintained in hot locations of the region, cooler sites offer potential for shifting towards new species and cultivars, especially for tree crops, which have particular climatic needs across seasons. Resilient options can be identified, but procedures to match tree species and cultivars with site climate need to be improved to facilitate effective adaptation.

There is ''high confidence'' that many oases of North Africa and the Arabian Peninsula are vulnerable to climate change. While the impacts of recent climate change are difficult to separate from the consequences of other change processes, it is ''likely'' that water resources have already declined in many places and the suitability of the local climate for many crops, especially perennial crops, has already decreased. This decline of water resources and thermal suitability of oasis locations for traditional crops is ''very likely'' to continue throughout the 21st century. In the coming years, the people living in oasis regions across the world will face challenges due to increasing impacts of global environmental change (Chen et al. 2018 <sup>[[#fn:r1719|1719]]</sup> ). Hence, efforts to increase their adaptive capacity to climate change can facilitate the sustainable development of oasis regions globally. In particular this will mean addressing the trade-offs between environmental restoration and agricultural livelihoods (Chen et al. 2018). Ultimately, sustainability in oasis regions will depend on policies integrating the provision of ecosystem services and social and human welfare needs (Wang et al. 2017 <sup>[[#fn:r1724|1724]]</sup> ).

<div id="section-3-7-4-oases-in-hyper-arid-areas-in-the-arabian-peninsula-and-northern-africa-block-7"></div>

<span id="figure-3.16"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.16'''

<span id="satellite-image-of-the-siwa-oasis-egypt.-source-google-maps"></span>
<!-- IMG CAPTION -->
'''Satellite image of the Siwa Oasis, Egypt. Source: Google Maps'''
<!-- IMG FILE -->
[[File:3b10dd13592196b01659391db6b2fec1 Figure-3.16-1024x961.jpg]]

Satellite image of the Siwa Oasis, Egypt. Source: Google Maps

<!-- END IMG -->
<span id="integrated-watershed-management"></span>
=== 3.7.5 Integrated watershed management ===

<div id="section-3-7-5-integrated-watershed-management-block-1"></div>

Desertification has resulted in significant loss of ecosystem processes and services, as described in detail in this chapter. The techniques and processes to restore degraded watersheds are not linear and integrated watershed management (IWM) must address physical, biological and social approaches to achieve SLM objectives (German et al. 2007 <sup>[[#fn:r1726|1726]]</sup> ).

<div id="section-3-7-5-1-jordan"></div>

<span id="jordan"></span>
==== 3.7.5.1 Jordan ====

<div id="section-3-7-5-1-jordan-block-1"></div>

Population growth, migration into Jordan and changes in climate have resulted in desertification of the Jordan Badia region. The Badia region covers more than 80% of the country’s area and receives less than 200 mm of rainfall per year, with some areas receiving less than 100 mm (Al-Tabini et al. 2012 <sup>[[#fn:r1727|1727]]</sup> ). Climate analysis has indicated a generally increasing dryness over the West Asia and Middle East region (AlSarmi and Washington 2011 <sup>[[#fn:r1728|1728]]</sup> ; Tanarhte et al. 2015 <sup>[[#fn:r1729|1729]]</sup> ), with reduction in average annual rainfall in Jordan’s Badia area (De Pauw et al. 2015 <sup>[[#fn:r1730|1730]]</sup> ). The incidence of extreme rainfall events has not declined over the region. Locally increased incidence of extreme events over the Mediterranean region has been proposed (Giannakopoulos et al. 2009 <sup>[[#fn:r1731|1731]]</sup> ).

The practice of intensive and localised livestock herding, in combination with deep ploughing and unproductive barley agriculture, are the main drivers of severe land degradation and depletion of the rangeland natural resources. This affected both the quantity and the diversity of vegetation as native plants with a high nutrition value were replaced with invasive species with low palatability and nutritional content (Abu-Zanat et al. 2004 <sup>[[#fn:r1732|1732]]</sup> ). The sparsely covered and crusted soils in Jordan’s Badia area have a low rainfall interception and infiltration rate, which leads to increased surface runoff and subsequent erosion and gullying, speeding up the drainage of rainwater from the watersheds, which can result in downstream flooding in Amman, Jordan (Oweis 2017 <sup>[[#fn:r1733|1733]]</sup> ).

<div id="section-3-7-5-1-jordan-block-2"></div>

<span id="figure-3.17a"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.17a'''

<span id="anewly-prepared-micro-water-harvesting-catchment-using-the-vallerani-system."></span>
<!-- IMG CAPTION -->
'''(a)Newly prepared micro water harvesting catchment, using the Vallerani system.'''
<!-- IMG FILE -->
[[File:1fb888a1be66840d4733d4e59b7d5d71 Figure-3.17a-e1575973543995-1024x576.jpg]]

(a)Newly prepared micro water harvesting catchment, using the Vallerani system.

<!-- END IMG -->
<div id="section-3-7-5-1-jordan-block-3"></div>

<span id="figure-3.17b"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure-3.17b'''

<span id="b-aerial-imaging-showing-micro-water-harvesting-catchment-treatment-after-planting"></span>
<!-- IMG CAPTION -->
'''(b) Aerial imaging showing micro water harvesting catchment treatment after planting'''
<!-- IMG FILE -->
[[File:c46f91b0be56e92c0e20ab15e4538313 Figure-3.17b-1024x768.jpg]]

(b) Aerial imaging showing micro water harvesting catchment treatment after planting

<!-- END IMG -->
<div id="section-3-7-5-1-jordan-block-4"></div>

<span id="figure-3.17c"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.17c'''

<span id="c-one-year-after-treatment.-source-stefan-strohmeier."></span>
<!-- IMG CAPTION -->
'''(c) one year after treatment. Source: Stefan Strohmeier.'''
<!-- IMG FILE -->
[[File:d5aac8ae1c063d1b3f3770d241841a5a Figure-3.17c.jpg]]

(c) one year after treatment. Source: Stefan Strohmeier.

<!-- END IMG -->
<div id="section-3-7-5-1-jordan-block-5"></div>

<span id="figure-3.18"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.18'''

<span id="illustration-of-enhanced-soil-water-retention-in-the-mechanized-micro-rainwater-harvesting-compared-to-untreated-badia-rangelands-in-jordan-showing-precipitation-pcp-sustained-stress-level-resulting-in-decreased-production-field-capacity-and-wilting-point-for-available-soil-moisture-and-then-measured-soil-moisture-content-between-the-two-treatments-degraded-rangeland-and-the-restored-rangeland-with"></span>
<!-- IMG CAPTION -->
'''Illustration of enhanced soil water retention in the Mechanized Micro Rainwater Harvesting compared to untreated Badia rangelands in Jordan, showing precipitation (PCP), sustained stress level resulting in decreased production, field capacity and wilting point for available soil moisture, and then measured soil moisture content between the two treatments (degraded rangeland and the restored rangeland with […]'''
<!-- IMG FILE -->
[[File:5855d365bf8274031e1d2d6a42eef563 Figure-3.18-1024x604.jpg]]

Illustration of enhanced soil water retention in the Mechanized Micro Rainwater Harvesting compared to untreated Badia rangelands in Jordan, showing precipitation (PCP), sustained stress level resulting in decreased production, field capacity and wilting point for available soil moisture, and then measured soil moisture content between the two treatments (degraded rangeland and the restored rangeland with the Vallerani plough).

<!-- END IMG -->
<div id="section-3-7-5-1-jordan-block-6"></div>

To restore the desertified Badia an IWM plan was developed using hillslope-implemented water harvesting micro catchments as a targeted restoration approach (Tabieh et al. 2015 <sup>[[#fn:r1734|1734]]</sup> ). Mechanized Micro Rainwater Harvesting (MIRWH) technology using the ‘Vallerani plough’ (Antinori and Vallerani 1994 <sup>[[#fn:r1735|1735]]</sup> ; Gammoh and Oweis 2011 <sup>[[#fn:r1736|1736]]</sup> ; Ngigi 2003 <sup>[[#fn:r1737|1737]]</sup> ) is being widely applied for rehabilitation of highly degraded rangeland areas in Jordan. A tractor digs out small water harvesting pits on the contour of the slope (Figure 3.17) allowing the retention, infiltration and local storage of surface runoff in the soil (Oweis 2017 <sup>[[#fn:r1739|1739]]</sup> ). The micro catchments are planted with native shrub seedlings, such as saltbush ( ''Atriplex halimus'' ), with enhanced survival as a function of increased soil moisture (Figure 3.18) and increased dry matter yields (&gt;300 kg ha <sup>–1</sup> ) that can serve as forage for livestock (Oweis 2017 <sup>[[#fn:r1738|1738]]</sup> ; Tabieh et al. 2015 <sup>[[#fn:r1740|1740]]</sup> ).

Simultaneously to MIRWH upland measures, the gully erosion is being treated through intermittent stone plug intervention (Figure 3.19), stabilising the gully beds, increasing soil moisture in proximity of the plugs, dissipating the surface runoff’s energy, and mitigating further back-cutting erosion and quick drainage of water. Eventually, the treated gully areas silt up and dense vegetation cover can re-establish. In addition, grazing management practices are implemented to increase the longevity of the treatment. Ultimately, the recruitment processes and re-vegetation shall control the watershed’s hydrological regime through rainfall interception, surface runoff deceleration and filtration, combined with the less erodible and enhanced infiltration characteristics of the rehabilitated soils.

<div id="section-3-7-5-1-jordan-block-7"></div>

<span id="figure-3.19a"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.19a'''

<span id="a-gully-plug-development-in-september-2017."></span>
<!-- IMG CAPTION -->
'''(a) Gully plug development in September 2017.'''
<!-- IMG FILE -->
[[File:f605e0a6fd89abf71f911bc3d3618186 Figure-3.19a.jpg]]

(a) Gully plug development in September 2017.

<!-- END IMG -->
<div id="section-3-7-5-1-jordan-block-8"></div>

<span id="figure-3.19b"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 3.19b'''

<span id="b-post-rainfall-event-march-2018.-near-amman-jordan.-source-stefan-strohmeier."></span>
<!-- IMG CAPTION -->
'''(b) Post-rainfall event (March 2018). Near Amman, Jordan. Source: Stefan Strohmeier.'''
<!-- IMG FILE -->
[[File:db037d1c28a821352a091a6f25dd9046 Figure-3.19b.jpg]]

(b) Post-rainfall event (March 2018). Near Amman, Jordan. Source: Stefan Strohmeier.

<!-- END IMG -->
<div id="section-3-7-5-1-jordan-block-9"></div>

In-depth understanding of the Badia’s rangeland status transition, coupled with sustainable rangeland management, are still subject to further investigation, development and adoption; a combination of all three is required to mitigate the ongoing degradation of the Middle Eastern rangeland ecosystems.

Oweis (2017) <sup>[[#fn:r1813|1813]]</sup> indicated that the cost of the fully automated Vallerani technique was approximately 32 USD ha-1. The total cost of the restoration package included the production, planting and maintenance of the shrub seedlings (11 USD ha <sup>–1)</sup> . Tabieh et al. (2015) <sup>[[#fn:r1812|1812]]</sup> calculated a benefit-cost ratio (BCR) of above 1.5 for re-vegetation of degraded Badia areas through MIRWH and saltbush. However, costs vary based on the seedling’s costs and availability of trained labour.

Water harvesting is not a recent scientific advancement. Water harvesting is known to have been developed during the Bronze Age and was widely practiced in the Negev Desert during the Byzantine time period (1300–1600 years ago) (Fried et al. 2018 <sup>[[#fn:r1741|1741]]</sup> ; Stavi et al. 2017 <sup>[[#fn:r1742|1742]]</sup> ). Through construction of various structures made of packed clay and stone, water was either held on site in half-circular dam structures ( ''hafir)'' that faced up-slope to capture runoff, or on terraces that slowed water allowing it to infiltrate and to be stored in the soil profile. Numerous other systems were designed to capture water in below-ground cisterns to be used later to provide water to livestock or for domestic use. Other water harvesting techniques divert runoff from hillslopes or wadis and spread the water in a systematic manner across ''playas'' and the toe-slope of a hillslope. These systems allow production of crops in areas with 100 mm of average annual precipitation by harvesting an additional 300+ mm of water (Beckers et al. 2013 <sup>[[#fn:r1743|1743]]</sup> ). Water harvesting is a proven technology to mitigate or adapt to climate change where precipitation may be reduced, and allow for small-scale crop and livestock production to continue supporting local needs.

<div id="section-3-7-5-2-india"></div>

<span id="india"></span>
==== 3.7.5.2 India ====

<div id="section-3-7-5-2-india-block-1"></div>

The second great challenge after the Green Revolution in India was the low productivity in the rain-fed and semi-arid regions where land degredation and drought were serious concerns. In response to this challenge IWM projects were implemented over large areas in semi-arid biomes over the past few decades. IWM was meant to become a key factor in meeting a range of social development goals in many semi-arid rainfed agrarian landscapes in India (Bouma et al. 2007 <sup>[[#fn:r1744|1744]]</sup> ; Kerr et al. 2002 <sup>[[#fn:r1745|1745]]</sup> ). Over the years, watershed development has become the fulcrum of rural development, and has the potential to achieve the twin objectives of ecosystem restoration and livelihood assurance in the drylands of India (Joy et al. 2004).

Many reports indicate significant improvements in mitigation of drought impacts, raising crops and fodder, livestock productivity, expanding the availability of drinking water and increasing incomes as a result of IWM (Rao 2000), but in some cases overall the positive impact of the programme has been questioned and, except in a few cases, the performance has not lived up to expectations (Joy et al. 2004; JM Kerr et al. 2002). Comparisons of catchments with and without IWM projects using remotely sensed data have sometimes shown no significant enhancement of biomass, in part due to methodological challenges of space for time comparisons (Bhalla et al. 2013 <sup>[[#fn:r1746|1746]]</sup> ). The factors contributing to the successful cases were found to include effective participation of stakeholders in management (Rao 2000; Ratna Reddy et al. 2004 <sup>[[#fn:r1747|1747]]</sup> ).

Attribution of success in soil and water conservation measures was confounded by inadequate monitoring of rainfall variability and lack of catchment hydrologic indicators (Bhalla et al. 2013 <sup>[[#fn:r1748|1748]]</sup> ). Social and economic trade-offs included bias of benefits to downstream crop producers at the expense of pastoralists, women and upstream communities. This biased distribution of IWM benefits could potentially be addressed by compensation for environmental services between communities (Kerr et al. 2002 <sup>[[#fn:r1749|1749]]</sup> ). The successes in some areas also led to increased demand for water, especially groundwater, since there has been no corresponding social regulation of water use after improvement in water regime (Samuel et al. 2007 <sup>[[#fn:r1750|1750]]</sup> ). Policies and management did not ensure water allocation to sectors with the highest social and economic benefits (Batchelor et al. 2003 <sup>[[#fn:r1751|1751]]</sup> ). Limited field evidence of the positive impacts of rainwater harvesting at the local scale is available, but there are several potential negative impacts at the watershed scale (Glendenning et al. 2012 <sup>[[#fn:r1752|1752]]</sup> ). Furthermore, watershed projects are known to have led to more water scarcity, and higher expectations for irrigation water supply, further exacerbating water scarcity (Bharucha et al. 2014 <sup>[[#fn:r1753|1753]]</sup> ).

In summary, the mixed performance of IWM projects has been linked to several factors. These include: inequity in the distribution of benefits (Kerr et al. 2002); focus on institutional aspects rather than application of appropriate watershed techniques and functional aspects of watershed restoration (Joy et al. 2006; Vaidyanathan 2006 <sup>[[#fn:r1755|1755]]</sup> ); mismatch between scales of focus and those that are optimal for catchment processes (Kerr 2007 <sup>[[#fn:r1756|1756]]</sup> ); inconsistencies in criteria used to select watersheds for IWM projects (Bhalla et al. 2011 <sup>[[#fn:r1757|1757]]</sup> ); and in a few cases additional costs and inefficiencies of local non-governmental organisations (Chandrasekhar et al. 2006 <sup>[[#fn:r1758|1758]]</sup> ; Deshpande 2008 <sup>[[#fn:r1759|1759]]</sup> ). Enabling policy responses for improvement of IWM performance include: a greater emphasis on ecological restoration rather than civil engineering; sharper focus on sustainability of livelihoods than just conservation; adoption of ‘water justice’ as a normative goal and minimising externalities on non-stakeholder communities; rigorous independent biophysical monitoring, with feedback mechanisms and integration with larger schemes for food and ecological security, and maintenance of environmental flows for downstream areas (Bharucha et al. 2014 <sup>[[#fn:r1760|1760]]</sup> ; Calder et al. 2008 <sup>[[#fn:r1761|1761]]</sup> ; Joy et al. 2006). Successful adaptation of IWM to achieve land degradation neutrality would largely depend on how IWM creatively engages with dynamics of large-scale land use and hydrology under a changing climate, involvement of livelihoods and rural incomes in ecological restoration, regulation of groundwater use, and changing aspirations of rural population ( ''robust evidence, high agreement'' ) (O’Brien et al. 2004 <sup>[[#fn:r1762|1762]]</sup> ; Samuel et al. 2007 <sup>[[#fn:r1763|1763]]</sup> ; Samuel and Joy 2018 <sup>[[#fn:r1764|1764]]</sup> ).

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==== 3.7.5.3 Limpopo River Basin ====

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Covering an area of 412,938 km <sup>2</sup> , the Limpopo River basin spans parts of Botswana, South Africa, Zimbabwe and Mozambique, eventually entering into the Mozambique Channel. It has been selected as a case study as it provides a clear illustration of the combined effect of desertification and climate change, and why IWM may be a crucial component of reducing exposure to climate change. It is predominantly a semi-arid area with an average annual rainfall of 400 mm (Mosase and Ahiablame 2018 <sup>[[#fn:r1765|1765]]</sup> ). Rainfall is both highly seasonal and variable, with the prominent impact of the El Niño/ La Niña phenomena and the Southern Oscillation leading to severe droughts (Jury 2016 <sup>[[#fn:r1766|1766]]</sup> ). It is also exposed to tropical cyclones that sweep in from the Mozambique Channel often leading to extensive casualties and the destruction of infrastructure (Christie and Hanlon 2001 <sup>[[#fn:r1767|1767]]</sup> ). Furthermore, there is good agreement across climate models that the region is going to become warmer and drier, with a change in the frequency of floods and droughts (Engelbrecht et al. 2011 <sup>[[#fn:r1768|1768]]</sup> ; Zhu and Ringler 2012). Seasonality is predicted to increase, which in turn may increase the frequency of flood events in an area that is already susceptible to flooding (Spaliviero et al. 2014 <sup>[[#fn:r1769|1769]]</sup> ).

A clear need exists to both address exposure to flood events as well as predicted decreases in water availability, which are already acute. Without the additional impact of climate change, the basin is rapidly reaching a point where all available water has been allocated to users (Kahinda et al. 2016 <sup>[[#fn:r1770|1770]]</sup> ; Zhu and Ringler 2012). The urgency of the situation was identified several decades ago (FAO 2004), with the countries of the basin recognising that responses are required at several levels, both in terms of system governance and the need to address land degradation.

Recent reviews of the governance and implementation of IWM within the basin recognise that an integrated approach is needed and that a robust institutional, legal, political, operational, technical and support environment is crucial (Alba et al. 2016 <sup>[[#fn:r1771|1771]]</sup> ; Gbetibouo et al. 2010 <sup>[[#fn:r1773|1773]]</sup> ; Machethe et al. 2004 <sup>[[#fn:r1774|1774]]</sup> ; Spaliviero et al. 2011 <sup>[[#fn:r1775|1775]]</sup> ; van der Zaag and Savenije 1999 <sup>[[#fn:r1776|1776]]</sup> ). Within the scope of emerging lessons, two principal ones emerge. The first is capacity and resource constraints at most levels. Limited capacity within Limpopo Watercourse Commission (LIMCOM) and national water management authorities constrains the implementation of IWM planning processes (Kahinda et al. 2016 <sup>[[#fn:r1777|1777]]</sup> ; Spaliviero et al. 2011 <sup>[[#fn:r1778|1778]]</sup> ). Whereas strategy development is often relatively well-funded and resourced through donor funding, long-term implementation is often limited due to competing priorities. The second is adequate representation of all parties in the process in order to address existing inequalities and ensure full integration of water management. For example, within Mozambique, significant strides have been made towards the decentralisation of river basin governance and IWM. Despite good progress, Alba et al. (2016) found that the newly implemented system may enforce existing inequalities as not all stakeholders, particularly smallholder farmers, are adequately represented in emerging water management structures and are often inhibited by financial and institutional constraints. Recognising economic and socio-political inequalities, and explicitly considering them to ensure the representation of all participants, can increase the chances of successful IWM implementation.

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