Editing IPCC:AR6/SRCCL/Chapter-4 (section)

== 4.9 Case studies ==

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Climate change impacts on land degradation can be avoided, reduced or even reversed, but need to be addressed in a context-sensitive manner. Many of the responses described in this section can also provide synergies of adaptation and mitigation. In this section we provide more in-depth analysis of a number of salient aspects of how land degradation and climate change interact. Table 4.3 is a synthesis of how of these case studies relate to climate change and other broader issues in terms of co-benefits.

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'''Table 4.3'''

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'''Synthesis of how the case studies interact with climate change and a broader set of co-benefits.'''
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[[File:bbd780f4dff0c0039b38655d00005f1d table-4.3.png]]

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=== 4.9.1 Urban green infrastructure ===

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Over half of the world’s population now lives in towns and cities, a proportion that is predicted to increase to about 70% by the middle of the century (United Nations 2015 <sup>[[#fn:r1226|1226]]</sup> ). Rapid urbanisation is a severe threat to land and the provision of ecosystem services (Seto et al. 2012 <sup>[[#fn:r1227|1227]]</sup> ). However, as cities expand, the avoidance of land degradation, or the maintenance/enhancement of ecosystem services is rarely considered in planning processes. Instead, economic development and the need for space for construction is prioritised, which can result in substantial pollution of air and water sources, the degradation of existing agricultural areas and indigenous, natural or semi-natural ecosystems both within and outside of urban areas. For instance, urban areas are characterised by extensive impervious surfaces. Degraded, sealed soils beneath these surfaces do not provide the same quality of water retention as intact soils. Urban landscapes comprising 50–90% impervious surfaces can therefore result in 40–83% of rainfall becoming surface water runoff (Pataki et al. 2011 <sup>[[#fn:r1228|1228]]</sup> ). With rainfall intensity predicted to increase in many parts of the world under climate change (Royal Society 2016 <sup>[[#fn:r1229|1229]]</sup> ), increased water runoff is going to get worse. Urbanisation, land degradation and climate change are therefore strongly interlinked, suggesting the need for common solutions (Reed and Stringer 2016 <sup>[[#fn:r1230|1230]]</sup> ).

There is now a large body of research and application demonstrating the importance of retaining urban green infrastructure (UGI) for the delivery of multiple ecosystem services (DG Environment News Alert Service, 2012; Wentworth, 2017 <sup>[[#fn:r1231|1231]]</sup> ) as an important tool to mitigate and adapt to climate change. UGI can be defined as all green elements within a city, including, but not limited to, retained indigenous ecosystems, parks, public greenspaces, green corridors, street trees, urban forests, urban agriculture, green roofs/walls and private domestic gardens (Tzoulas et al. 2007 <sup>[[#fn:r1232|1232]]</sup> ). The definition is usually extended to include ‘blue’ infrastructure, such as rivers, lakes, bioswales and other water drainage features. The related concept of Nature-Based Solutions (defined as: ''living solutions inspired by, continuously supported by and using nature, which are designed to address various societal challenges in a resource-efficient and adaptable manner and to provide simultaneously economic, social, and environmental benefits'' ) has gained considerable traction within the European Commission as one approach to mainstreaming the importance of UGI (Maes and Jacobs 2017 <sup>[[#fn:r1233|1233]]</sup> ; European Union 2015 <sup>[[#fn:r1234|1234]]</sup> ).

Through retaining existing vegetation and ecosystems, revegetating previous developed land or integrating vegetation into buildings in the form of green walls and roofs, UGI can play a direct role in mitigating climate change through carbon sequestration. However, compared to overall carbon emissions from cities, effects will be small. Given that UGI necessarily involves the retention and management of non-sealed surfaces, co-benefits for land degradation (e.g., soil compaction avoidance, reduced water runoff, carbon storage and vegetation productivity (Davies et al. 2011 <sup>[[#fn:r1235|1235]]</sup> ; Edmondson et al. 2011 <sup>[[#fn:r1236|1236]]</sup> , 2014 <sup>[[#fn:r1237|1237]]</sup> ; Yao et al. 2015 <sup>[[#fn:r1238|1238]]</sup> ) will also be apparent. Although not currently a priority, its role in mitigating land degradation could be substantial. For instance, appropriately managed innovative urban agricultural production systems, such as vertical farms, could have the potential to meet some of the food needs of cities and reduce the production (and therefore degradation) pressure on agricultural land in rural areas, although thus far this is unproven (for a recent review, see Wilhelm and Smith 2018).

The importance of UGI as part of a climate change adaptation approach has received greater attention and application (Gill et al. 2007 <sup>[[#fn:r1239|1239]]</sup> ; Fryd et al. 2011 <sup>[[#fn:r1240|1240]]</sup> ; Demuzere et al. 2014 <sup>[[#fn:r1241|1241]]</sup> ; Sussams et al. 2015 <sup>[[#fn:r1242|1242]]</sup> ). The EU’s Adapting to Climate Change white paper emphasises the ‘crucial role in adaptation in providing essential resources for social and economic purposes under extreme climate conditions’ (CEC, 2009, p. 9). Increasing vegetation cover, planting street trees and maintaining/expanding public parks reduces temperatures (Cavan et al. 2014 <sup>[[#fn:r1243|1243]]</sup> ; Di Leo et al. 2016 <sup>[[#fn:r1244|1244]]</sup> ; Feyisa et al. 2014 <sup>[[#fn:r1245|1245]]</sup> ; Tonosaki and Kawai 2014 <sup>[[#fn:r1246|1246]]</sup> ; Zölch et al. 2016 <sup>[[#fn:r1247|1247]]</sup> ). Further, the appropriate design and spatial distribution of greenspaces within cities can help to alter urban climates to improve human health and comfort (e.g., Brown and Nicholls 2015 <sup>[[#fn:r1248|1248]]</sup> ; Klemm et al. 2015 <sup>[[#fn:r1249|1249]]</sup> ). The use of green walls and roofs can also reduce energy use in buildings (e.g., Coma et al. 2017 <sup>[[#fn:r1250|1250]]</sup> ). Similarly, natural flood management and ecosystem-based approaches of providing space for water, renaturalising rivers and reducing surface runoff through the presence of permeable surfaces and vegetated features (including walls and roofs) can manage flood risks, impacts and vulnerability (e.g., Gill et al. 2007 <sup>[[#fn:r1251|1251]]</sup> ; Munang et al. 2013 <sup>[[#fn:r1252|1252]]</sup> ). Access to UGI in times of environmental stresses and shock can provide safety nets for people, and so can be an important adaptation mechanism, both to climate change (Potschin et al. 2016 <sup>[[#fn:r1253|1253]]</sup> ) and land degradation.

Most examples of UGI implementation as a climate change adaptation strategy have centred on its role in water management for flood risk reduction. The importance for land degradation is either not stated, or not prioritised. In Beira, Mozambique, the government is using UGI to mitigate increased flood risks predicted to occur under climate change and urbanisation, which will be done by improving the natural water capacity of the Chiveve River. As part of the UGI approach, mangrove habitats have been restored, and future phases include developing new multi-functional urban green spaces along the river (World Bank 2016 <sup>[[#fn:r1254|1254]]</sup> ). The retention of green spaces within the city will have the added benefit of halting further degradation in those areas. Elsewhere, planning mechanisms promote the retention and expansion of green areas within cities to ensure ecosystem service delivery, which directly halts land degradation, but are largely viewed and justified in the context of climate change adaptation and mitigation. For instance, the Berlin Landscape Programme includes five plans, one of which covers adapting to climate change through the recognition of the role of UGI (Green Surge 2016 <sup>[[#fn:r1255|1255]]</sup> ). Major climate-related challenges facing Durban, South Africa, include sea level rise, urban heat island, water runoff and conservation (Roberts and O’Donoghue 2013 <sup>[[#fn:r1256|1256]]</sup> ). Now considered a global leader in climate adaptation planning (Roberts 2010 <sup>[[#fn:r1257|1257]]</sup> ), Durban’s Climate Change Adaptation plan includes the retention and maintenance of natural ecosystems, in particular those that are important for mitigating flooding, coastal erosion, water pollution, wetland siltation and climate change (eThekwini Municipal Council 2014 <sup>[[#fn:r1258|1258]]</sup> ).

<span id="perennial-grains-and-soil-organic-carbon-soc"></span>
=== 4.9.2 Perennial grains and soil organic carbon (SOC) ===

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The severe ecological perturbation that is inherent in the conversion of native perennial vegetation to annual crops, and the subsequent high frequency of perturbation required to maintain annual crops, results in at least four forms of soil degradation that will be exacerbated by the effects of climate change (Crews et al. 2016 <sup>[[#fn:r1259|1259]]</sup> ). First, soil erosion is a very serious consequence of annual cropping, with median losses exceeding rates of formation by one to two orders of magnitude in conventionally plowed agroecosystems, and while erosion is reduced with conservation tillage, median losses still exceed formation by several fold (Montgomery 2007 <sup>[[#fn:r1260|1260]]</sup> ). More severe storm intensity associated with climate change is expected to cause even greater losses to wind and water erosion (Nearing et al. 2004 <sup>[[#fn:r1261|1261]]</sup> ). Second, the periods of time in which live roots are reduced or altogether absent from soils in annual cropping systems allow for substantial losses of nitrogen from fertilised croplands, averaging 50% globally (Ladha et al. 2005 <sup>[[#fn:r1262|1262]]</sup> ). This low retention of nitrogen is also expected to worsen with more intense weather events (Bowles et al. 2018 <sup>[[#fn:r1263|1263]]</sup> ). A third impact of annual cropping is the degradation of soil structure caused by tillage, which can reduce infiltration of precipitation, and increase surface runoff. It is predicted that the percentage of precipitation that infiltrates into agricultural soils will decrease further under climate-change scenarios (Basche and DeLonge 2017 <sup>[[#fn:r1264|1264]]</sup> ; Wuest et al. 2006 <sup>[[#fn:r1265|1265]]</sup> ). The fourth form of soil degradation that results from annual cropping is the reduction of soil organic matter (SOM), a topic of particular relevance to climate change mitigation and adaptation.

Undegraded cropland soils can theoretically hold far more SOM (which is about 58% carbon) than they currently do (Soussana et al. 2006 <sup>[[#fn:r1266|1266]]</sup> ). We know this deficiency because, with few exceptions, comparisons between cropland soils and those of proximate mature native ecosystems commonly show a 40–75% decline in soil carbon attributable to agricultural practices. What happens when native ecosystems are converted to agriculture that induces such significant losses of SOM? Wind and water erosion commonly results in preferential removal of light organic matter fractions that can accumulate on or near the soil surface (Lal 2003 <sup>[[#fn:r1267|1267]]</sup> ). In addition to the effects of erosion, the fundamental practices of growing annual food and fibre crops alters both inputs and outputs of organic matter from most agroecosystems, resulting in net reductions in soil carbon equilibria (Soussana et al. 2006 <sup>[[#fn:r1268|1268]]</sup> ; McLauchlan 2006 <sup>[[#fn:r1269|1269]]</sup> ; Crews et al. 2016 <sup>[[#fn:r1270|1270]]</sup> ). Native vegetation of almost all terrestrial ecosystems is dominated by perennial plants, and the below-ground carbon allocation of these perennials is a key variable in determining formation rates of stable soil organic carbon (SOC) (Jastrow et al. 2007 <sup>[[#fn:r1271|1271]]</sup> ; Schmidt et al. 2011 <sup>[[#fn:r1272|1272]]</sup> ). When perennial vegetation is replaced by annual crops, inputs of root-associated carbon (roots, exudates, mycorrhizae) decline substantially. For example, perennial grassland species allocate around 67% of productivity to roots, whereas annual crops allocate between 13–30% (Saugier 2001 <sup>[[#fn:r1273|1273]]</sup> ; Johnson et al. 2006 <sup>[[#fn:r1274|1274]]</sup> ).

At the same time, inputs of SOC are reduced in annual cropping systems, and losses are increased because of tillage, compared to native perennial vegetation. Tillage breaks apart soil aggregates which, among other functions, are thought to inhibit soil bacteria, fungi and other microbes from consuming and decomposing SOM (Grandy and Neff 2008 <sup>[[#fn:r1275|1275]]</sup> ). Aggregates reduce microbial access to organic matter by restricting physical access to mineral-stabilised organic compounds as well as reducing oxygen availability (Cotrufo et al. 2015 <sup>[[#fn:r1276|1276]]</sup> ; Lehmann and Kleber 2015 <sup>[[#fn:r1277|1277]]</sup> ). When soil aggregates are broken open with tillage in the conversion of native ecosystems to agriculture, microbial consumption of SOC and subsequent respiration of CO <sub>2</sub> increase dramatically, reducing soil carbon stocks (Grandy and Robertson 2006 <sup>[[#fn:r1278|1278]]</sup> ; Grandy and Neff 2008 <sup>[[#fn:r1279|1279]]</sup> ).

Many management approaches are being evaluated to reduce soil degradation in general, especially by increasing mineral-protected forms of SOC in the world’s croplands (Paustian et al. 2016 <sup>[[#fn:r1280|1280]]</sup> ). The menu of approaches being investigated focuses either on increasing below-ground carbon inputs, usually through increases in total crop productivity, or by decreasing microbial activity, usually through reduced soil disturbance (Crews and Rumsey 2017 <sup>[[#fn:r1281|1281]]</sup> ). However, the basic biogeochemistry of terrestrial ecosystems managed for production of annual crops presents serious challenges to achieving the standing stocks of SOC accumulated by native ecosystems that preceded agriculture. A novel new approach that is just starting to receive significant attention is the development of perennial cereal, legume and oilseed crops (Glover et al. 2010 <sup>[[#fn:r1282|1282]]</sup> ; Baker 2017 <sup>[[#fn:r1283|1283]]</sup> ).

There are two basic strategies that plant breeders and geneticists are using to develop new perennial grain crop species. The first involves making wide hybrid crosses between existing elite lines of annual crops, such as wheat, sorghum and rice, with related wild perennial species in order to introgress perennialism into the genome of the annual (Cox et al. 2018 <sup>[[#fn:r1284|1284]]</sup> ; Huang et al. 2018 <sup>[[#fn:r1285|1285]]</sup> ; Hayes et al. 2018 <sup>[[#fn:r1286|1286]]</sup> ). The other approach is de ''novo'' domestication of wild perennial species that have crop-like traits of interest (DeHaan et al. 2016 <sup>[[#fn:r1287|1287]]</sup> ; DeHaan and Van Tassel 2014 <sup>[[#fn:r1288|1288]]</sup> ). New perennial crop species undergoing de ''novo'' domestication include intermediate wheatgrass, a relative of wheat that produces grain known as Kernza (DeHaan et al. 2018 <sup>[[#fn:r1289|1289]]</sup> ; Cattani and Asselin 2018 <sup>[[#fn:r1290|1290]]</sup> ) and ''Silphium integrifolium'' , an oilseed crop in the sunflower family (Van Tassel et al. 2017 <sup>[[#fn:r1291|1291]]</sup> ). Other grain crops receiving attention for perennialisation include pigeon pea, barley, buckwheat and maize (Batello et al. 2014 <sup>[[#fn:r1292|1292]]</sup> ; Chen et al. 2018c <sup>[[#fn:r1293|1293]]</sup> ) and a number of legume species (Schlautman et al. 2018 <sup>[[#fn:r1294|1294]]</sup> ). In most cases, the seed yields of perennial grain crops under development are well below those of elite modern grain varieties. During the period that it will take for intensive breeding efforts to close the yield and other trait gaps between annual and perennial grains, perennial proto-crops may be used for purposes other than grain, including forage production (Ryan et al. 2018 <sup>[[#fn:r1295|1295]]</sup> ). Perennial rice stands out as a high-yielding exception, as its yields matched those of elite local varieties in the Yunnan Province for six growing seasons over three years (Huang et al. 2018 <sup>[[#fn:r1296|1296]]</sup> ).

In a perennial agroecosystem, the biogeochemical controls on SOC accumulation shift dramatically, and begin to resemble the controls that govern native ecosystems (Crews et al. 2016 <sup>[[#fn:r1297|1297]]</sup> ). When erosion is reduced or halted, and crop allocation to roots increases by 100–200%, and when soil aggregates are not disturbed thus reducing microbial respiration, SOC levels are expected to increase (Crews and Rumsey 2017 <sup>[[#fn:r1298|1298]]</sup> ). Deep roots growing year round are also effective at increasing nitrogen retention (Culman et al. 2013 <sup>[[#fn:r1299|1299]]</sup> ; Jungers et al. 2019 <sup>[[#fn:r1300|1300]]</sup> ). Substantial increases in SOC have been measured where croplands that had historically been planted to annual grains were converted to perennial grasses, such as in the US Conservation Reserve Program or in plantings of second-generation perennial biofuel crops. Two studies have assessed carbon accumulation in soils when croplands were converted to the perennial grain Kernza. In one, researchers found no differences in soil labile (permanganate-oxidisable) carbon after four years of cropping to perennial Kernza versus annual wheat in a sandy textured soil. Given that coarse textured soils do not offer the same physicochemical protection against microbial attack as many finer textured soils, these results are not surprising, but these results do underscore how variable the rates of carbon accumulation can be (Jastrow et al. 2007 <sup>[[#fn:r1301|1301]]</sup> ). In the second study, researchers assessed the carbon balance of a Kernza field in Kansas, USA over 4.5 years using eddy covariance observations (de Oliveira et al. 2018). They found that the net carbon accumulation rate of about 1500 gC m <sup>–2</sup> yr <sup>–1</sup> in the first year of the study corresponding to the biomass of Kernza, increasing to about 300 gC m <sup>–2</sup> yr <sup>–1</sup> in the final year, where CO <sub>2</sub> respiration losses from the decomposition of roots and SOM approached new carbon inputs from photosynthesis. Based on measurements of soil carbon accumulation in restored grasslands in this part of the USA, the net carbon accumulation in stable organic matter under a perennial grain crop might be expected to sequester 30–50 gC m <sup>–2</sup> yr <sup>–1</sup> (Post and Kwon 2000 <sup>[[#fn:r1302|1302]]</sup> ) until a new equilibrium is reached. Sugar cane, a highly productive perennial, has been shown to accumulate a mean of 187 gC m–2 yr <sup>–1</sup> in Brazil (La Scala Júnior et al. 2012 <sup>[[#fn:r1303|1303]]</sup> ).

Reduced soil erosion, increased nitrogen retention, greater water uptake efficiency and enhanced carbon sequestration represent improved ecosystem functions, made possible in part by deep and extensive root systems of perennial crops (Figure 4.8).

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'''Figure-4.8'''

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'''Comparison of root systems between the newly domesticated intermediate wheatgrass (left) and annual wheat (right). Photo: Copyright © Jim Richardson.'''
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[[File:377ce2965420e3fe2503b2279fee43e8 Figure-4.8-1024x683.jpg]]

Comparison of root systems between the newly domesticated intermediate wheatgrass (left) and annual wheat (right). Photo: Copyright © Jim Richardson.

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When compared to annual grains like wheat, single species stands of deep-rooted perennial grains such as Kernza are expected to reduce soil erosion, increase nitrogen retention, achieve greater water uptake efficiency and enhance carbon sequestration (Crews et al. 2018 <sup>[[#fn:r1304|1304]]</sup> ) (Figure 4.8). An even higher degree of ecosystem services can, at least theoretically, be achieved by strategically combining different functional groups of crops such as a cereal and a nitrogen-fixing legume (Soussana and Lemaire 2014 <sup>[[#fn:r1305|1305]]</sup> ). Not only is there evidence from plant-diversity experiments that communities with higher species richness sustain higher concentrations of SOC (Hungate et al. 2017 <sup>[[#fn:r1306|1306]]</sup> ; Sprunger and Robertson 2018 <sup>[[#fn:r1307|1307]]</sup> ; Chen, S. 2018 <sup>[[#fn:r1308|1308]]</sup> ; Yang et al. 2019 <sup>[[#fn:r1309|1309]]</sup> ), but other valuable ecosystem services such as pest suppression, lower GHG emissions, and greater nutrient retention may be enhanced (Schnitzer et al. 2011 <sup>[[#fn:r1310|1310]]</sup> ; Culman et al. 2013 <sup>[[#fn:r1311|1311]]</sup> ).

Similar to perennial forage crops such as alfalfa, perennial grain crops are expected to have a definite productive lifespan, probably in the range of three to 10 years. A key area of research on perennial grains cropping systems is to minimise losses of SOC during conversion of one stand of perennial grains to another. Recent work demonstrates that no-till conversion of a mature perennial grassland to another perennial crop will experience several years of high net CO <sub>2</sub> emissions as decomposition of copious crop residues exceed ecosystem uptake of carbon by the new crop (Abraha et al. 2018 <sup>[[#fn:r1312|1312]]</sup> ). Most, if not all, of this lost carbon will be recaptured in the replacement crop. It is not known whether mineral-stabilised carbon that is protected in soil aggregates is vulnerable to loss in perennial crop succession.

Perennial grains hold promises of agricultural practices, which can significantly reduce soil erosion and nutrient leakage while sequestering carbon. When cultivated in mixes with N-fixing species (legumes) such polycultures also reduce the need for external inputs of nitrogen – a large source of GHG from conventional agriculture.

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=== 4.9.3 Reversing land degradation through reforestation ===

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==== 4.9.3.1 South Korea case study on reforestation success ====

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In the first half of the 20th century, forests in the Republic of Korea (South Korea) were severely degraded and deforested during foreign occupations and the Korean War. Unsustainable harvest for timber and fuelwood resulted in severely degraded landscapes, heavy soil erosion and large areas denuded of vegetation cover. Recognising that South Korea’s economic health would depend on a healthy environment, South Korea established a national forest service (1967) and embarked on the first phase of a 10-year reforestation programme in 1973 (Forest Development Program), which was followed by subsequent reforestation programmes that ended in 1987, after 2.4 Mha of forests were restored (Figure 4.9).

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'''Figure 4.9'''

<span id="example-of-severely-degraded-hills-in-south-korea-and-stages-of-forest-restoration.-the-top-two-photos-are-taken-in-the-early-1970s-before-and-after-restoration-the-third-photo-about-five-years-after-restoration-and-the-bottom-photo-was-taken-about-20-years-after-restoration.-many-examples-of-such-restoration-success-exist-throughout-south"></span>
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'''Example of severely degraded hills in South Korea and stages of forest restoration. The top two photos are taken in the early 1970s, before and after restoration, the third photo about five years after restoration, and the bottom photo was taken about 20 years after restoration. Many examples of such restoration success exist throughout South […]'''
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[[File:fdb023a3ba1655c460f57efa8a7abd4e Figure-4.9-1024x576.jpg]]

Example of severely degraded hills in South Korea and stages of forest restoration. The top two photos are taken in the early 1970s, before and after restoration, the third photo about five years after restoration, and the bottom photo was taken about 20 years after restoration. Many examples of such restoration success exist throughout South Korea. (Photos: Copyright © Korea Forest Service)

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As a consequence of reforestation, forest volume increased from 11.3 m <sup>3</sup> ha–1 in 1973 to 125.6 m <sup>3</sup> ha–1 in 2010 and 150.2 m <sup>3</sup> ha–1 in 2016 (Korea Forest Service 2017 <sup>[[#fn:r1313|1313]]</sup> ). Increases in forest volume had significant co-benefits such as increasing water yield by 43% and reducing soil losses by 87% from 1971 to 2010 (Kim et al. 2017 <sup>[[#fn:r1314|1314]]</sup> ).

The forest carbon density in South Korea has increased from 5–7 MgC ha–1 in the period 1955–1973 to more than 30 MgC ha <sup>–1</sup> in the late 1990s (Choi et al. 2002 <sup>[[#fn:r1315|1315]]</sup> ). Estimates of carbon uptake rates in the late 1990s were 12 TgC yr <sup>–1</sup> (Choi et al. 2002 <sup>[[#fn:r1316|1316]]</sup> ). For the period 1954 to 2012, carbon uptake was 8.3 TgC yr <sup>–1</sup> (Lee et al. 2014 <sup>[[#fn:r1317|1317]]</sup> ), lower than other estimates because reforestation programmes did not start until 1973. Net ecosystem production in South Korea was 10.55 ± 1.09 TgC yr <sup>−1</sup> in the 1980s, 10.47 ± 7.28 Tg C yr <sup>−1</sup> in the 1990s, and 6.32 ± 5.02 Tg C yr <sup>−1</sup> in the 2000s, showing a gradual decline as average forest age increased (Cui et al. 2014 <sup>[[#fn:r1318|1318]]</sup> ). The estimated past and projected future increase in the carbon content of South Korea’s forest area during 1992–2034 was 11.8 TgC yr <sup>–1</sup> (Kim et al. 2016 <sup>[[#fn:r1319|1319]]</sup> ).

During the period of forest restoration, South Korea also promoted inter-agency cooperation and coordination, especially between the energy and forest sectors, to replace firewood with fossil fuels, and to reduce demand for firewood to help forest recovery (Bae et al. 2012 <sup>[[#fn:r1320|1320]]</sup> ). As experience with forest restoration programmes has increased, emphasis has shifted from fuelwood plantations, often with exotic species and hybrid varieties to planting more native species and encouraging natural regeneration (Kim and Zsuffa 1994 <sup>[[#fn:r1321|1321]]</sup> ; Lee et al. 2015 <sup>[[#fn:r1322|1322]]</sup> ). Avoiding monocultures in reforestation programmes can reduce susceptibility to pests (Kim and Zsuffa 1994 <sup>[[#fn:r1323|1323]]</sup> ). Other important factors in the success of the reforestation programme were that private landowners were heavily involved in initial efforts (both corporate entities and smallholders) and that the reforestation programme was made part of the national economic development programme (Lamb 2014 <sup>[[#fn:r1324|1324]]</sup> ).

The net present value and the cost-benefit ratio of the reforestation programme were 54.3 billion and 5.84 billion USD in 2010, respectively. The breakeven point of the reforestation investment appeared within two decades. Substantial benefits of the reforestation programme included disaster risk reduction and carbon sequestration (Lee et al. 2018a <sup>[[#fn:r1325|1325]]</sup> ).

In summary, the reforestation programme was a comprehensive technical and social initiative that restored forest ecosystems, enhanced the economic performance of rural regions, contributed to disaster risk reduction, and enhanced carbon sequestration (Kim et al. 2017 <sup>[[#fn:r1326|1326]]</sup> ; Lee et al. 2018a <sup>[[#fn:r1327|1327]]</sup> ; UNDP 2017 <sup>[[#fn:r1328|1328]]</sup> ).

The success of the reforestation programme in South Korea and the associated significant carbon sink indicate a high mitigation potential that might be contributed by a potential future reforestation programme in the Democratic People’s Republic of Korea (North Korea) (Lee et al. 2018b <sup>[[#fn:r1329|1329]]</sup> ).

<div id="section-4-9-3-2-china-case-study-on-reforestation-success"></div>

<span id="china-case-study-on-reforestation-success"></span>
==== 4.9.3.2 China case study on reforestation success ====

<div id="section-4-9-3-2-china-case-study-on-reforestation-success-block-1"></div>

The dramatic decline in the quantity and quality of natural forests in China resulted in land degradation, such as soil erosion, floods, droughts, carbon emission, and damage to wildlife habitat (Liu and Diamond 2008 <sup>[[#fn:r1330|1330]]</sup> ). In response to failures of previous forestry and land policies, the severe droughts in 1997, and the massive floods in 1998, the central government decided to implement a series of land degradation control policies, including the National Forest Protection Program (NFPP), Grain for Green or the Conversion of Cropland to Forests and Grassland Program (GFGP) (Liu et al. 2008 <sup>[[#fn:r1331|1331]]</sup> ; Yin 2009 <sup>[[#fn:r1332|1332]]</sup> ; Tengberg et al. 2016 <sup>[[#fn:r1333|1333]]</sup> ; Zhang et al. 2000 <sup>[[#fn:r1334|1334]]</sup> ). The NFPP aimed to completely ban logging of natural forests in the upper reaches of the Yangtze and Yellow rivers as well as in Hainan Province by 2000 and to substantially reduce logging in other places (Xu et al. 2006 <sup>[[#fn:r1335|1335]]</sup> ). In 2011, NFPP was renewed for the 10-year second phase, which also added another 11 counties around Danjiangkou Reservoir in Hubei and Henan Provinces, the water source for the middle route of the South-to-North Water Diversion Project (Liu et al. 2013 <sup>[[#fn:r1336|1336]]</sup> ). Furthermore, the NFPP afforested 31 Mha by 2010 through aerial seeding, artificial planting, and mountain closure (i.e., prohibition of human activities such as fuelwood collection and lifestock grazing) (Xu et al. 2006 <sup>[[#fn:r1337|1337]]</sup> ). China banned commercial logging in all natural forests by the end of 2016, which imposed logging bans and harvesting reductions in 68.2 Mha of forest land – including 56.4 Mha of natural forest (approximately 53% of China’s total natural forests).

GFGP became the most ambitious of China’s ecological restoration efforts, with more than 45 billion USD devoted to its implementation since 1990 (Kolinjivadi and Sunderland 2012 <sup>[[#fn:r1338|1338]]</sup> ) The programme involves the conversion of farmland on slopes of 15–25° or greater to forest or grassland (Bennett 2008 <sup>[[#fn:r1339|1339]]</sup> ). The pilot programme started in three provinces – Sichuan, Shaanxi and Gansu – in 1999 (Liu and Diamond 2008 <sup>[[#fn:r1340|1340]]</sup> ). After its initial success, it was extended to 17 provinces by 2000 and finally to all provinces by 2002, including the headwaters of the Yangtze and Yellow rivers (Liu et al. 2008).

NFPP and GFGP have dramatically improved China’s land conditions and ecosystem services, and thus have mitigated the unprecedented land degradation in China (Liu et al. 2013 <sup>[[#fn:r1341|1341]]</sup> ; Liu et al. 2002 <sup>[[#fn:r1342|1342]]</sup> ; Long et al. 2006 <sup>[[#fn:r1343|1343]]</sup> ; Xu et al. 2006 <sup>[[#fn:r1344|1344]]</sup> ). NFPP protected 107 Mha forest area and increased forest area by 10 Mha between 2000 and 2010. For the second phase (2011–2020), the NFPP plans to increase forest cover by a further 5.2 Mha, capture 416 million tons of carbon, provide 648,500 forestry jobs, further reduce land degradation, and enhance biodiversity (Liu et al. 2013 <sup>[[#fn:r1345|1345]]</sup> ). During 2000–2007, sediment concentration in the Yellow River had declined by 38%. In the Yellow River basin, it was estimated that surface runoff would be reduced by 450 million m3 from 2000 to 2020, which is equivalent to 0.76% of the total surface water resources (Jia et al. 2006). GFGP had cumulatively increased vegetative cover by 25 Mha, with 8.8 Mha of cropland being converted to forest and grassland, 14.3 Mha barren land being afforested, and 2.0 Mha of forest regeneration from mountain closure. Forest cover within the GFGP region has increased 2% during the first eight years (Liu et al. 2008 <sup>[[#fn:r1346|1346]]</sup> ). In Guizhou Province, GFGP plots had 35–53% less loss of phosphorus than non-GFGP plots (Liu et al. 2002 <sup>[[#fn:r1347|1347]]</sup> ). In Wuqi County of Shaanxi Province, the Chaigou Watershed had 48% and 55% higher soil moisture and moisture-holding capacity in GFGP plots than in non-GFGP plots, respectively (Liu et al. 2002 <sup>[[#fn:r1348|1348]]</sup> ). According to reports on China’s first national ecosystem assessment (2000–2010), for carbon sequestration and soil retention, coefficients for the GFGP targeting forest restoration and NFPP are positive and statistically significant. For sand fixation, GFGP targeting grassland restoration is positive and statistically significant. Remote sensing observations confirm that vegetation cover increased and bare soil declined in China over the period 2001 to 2015 (Qiu et al. 2017 <sup>[[#fn:r1349|1349]]</sup> ). But, where afforestation is sustained by drip irrigation from groundwater, questions about plantation sustainability arise (Chen et al. 2018a <sup>[[#fn:r1350|1350]]</sup> ). Moreover, greater gains in biodiversity could be achieved by promoting mixed forests over monocultures (Hua et al. 2016 <sup>[[#fn:r1351|1351]]</sup> ).

NFPP-related activities received a total commitment of 93.7 billion yuan (about 14 billion USD at 2018 exchange rate) between 1998 and 2009. Most of the money was used to offset economic losses of forest enterprises caused by the transformation from logging to tree plantations and forest management (Liu et al. 2008 <sup>[[#fn:r1352|1352]]</sup> ). By 2009, the cumulative total investment through the NFPP and GFGP exceeded 50 billion USD2009 and directly involved more than 120 million farmers in 32 million households in the GFGP alone (Liu et al. 2013 <sup>[[#fn:r1353|1353]]</sup> ). All programmes reduce or reverse land degradation and improve human well-being. Thus, a coupled human and natural systems perspective (Liu et al. 2008 <sup>[[#fn:r1354|1354]]</sup> ) would be helpful to understand the complexity of policies and their impacts, and to establish long-term management mechanisms to improve the livelihood of participants in these programmes and other land management policies in China and many other parts of the world.

<span id="degradation-and-management-of-peat-soils"></span>
=== 4.9.4 Degradation and management of peat soils ===

<div id="section-4-9-4-degradation-and-management-of-peat-soils-block-1"></div>

Globally, peatlands cover 3–4% of the Earth’s land area (about 430 Mha) (Xu et al. 2018a <sup>[[#fn:r1355|1355]]</sup> ) and store 26–44% of estimated global SOC (Moore 2002 <sup>[[#fn:r1356|1356]]</sup> ). They are most abundant in high northern latitudes, covering large areas in North America, Russia and Europe. At lower latitudes, the largest areas of tropical peatlands are located in Indonesia, the Congo Basin and the Amazon Basin in the form of peat swamp forests (Gumbricht et al. 2017 <sup>[[#fn:r1357|1357]]</sup> ; Xu et al. 2018a <sup>[[#fn:r1358|1358]]</sup> ). It is estimated that, while 80–85% of the global peatland areas is still largely in a natural state, they are such carbon-dense ecosystems that degraded peatlands (0.3% of the terrestrial land) are responsible for a disproportional 5% of global anthropogenic CO <sub>2</sub> emissions – that is, an annual addition of 0.9–3 GtCO <sub>2</sub> to the atmosphere (Dommain et al. 2012 <sup>[[#fn:r1359|1359]]</sup> ; IPCC 2014c <sup>[[#fn:r1360|1360]]</sup> ).

Peatland degradation is not well quantified globally, but regionally peatland degradation can involve a large percentage of the areas. Land-use change and degradation in tropical peatlands have primarily been quantified in Southeast Asia, where drainage and conversion to plantation crops is the dominant transition (Miettinen et al. 2016 <sup>[[#fn:r1361|1361]]</sup> ). Degradation of peat swamps in Peru is also a growing concern and one pilot survey showed that more than 70% of the peat swamps were degraded in one region surveyed (Hergoualc’h et al. 2017a <sup>[[#fn:r1362|1362]]</sup> ). Around 65,000 km2 or 10% of the European peatland area has been lost and 44% of the remaining European peatlands are degraded (Joosten, H., Tanneberger 2017 <sup>[[#fn:r1363|1363]]</sup> ). Large areas of fens have been entirely ‘lost’ or greatly reduced in thickness due to peat wastage (Lamers et al. 2015 <sup>[[#fn:r1364|1364]]</sup> ).

The main drivers of the acceleration of peatland degradation in the 20th century were associated with drainage for agriculture, peat extraction and afforestation related activities (burning, over-grazing, fertilisation) with a variable scale and severity of impact depending on existing resources in the various countries (O’Driscoll et al. 2018 <sup>[[#fn:r1365|1365]]</sup> ; Cobb, A.R. et al. Dommain et al. 2018 <sup>[[#fn:r1366|1366]]</sup> ; Lamers et al. 2015 <sup>[[#fn:r1367|1367]]</sup> ). New drivers include urban development, wind farm construction (Smith et al. 2012 <sup>[[#fn:r1368|1368]]</sup> ), hydroelectric development, tar sands mining and recreational uses (Joosten and Tanneberger 2017 <sup>[[#fn:r1369|1369]]</sup> ). Anthropogenic pressures are now affecting peatlands in previously geographically isolated areas with consequences for global environmental concerns and impacts on local livelihoods (Dargie et al. 2017 <sup>[[#fn:r1370|1370]]</sup> ; Lawson et al. 2015 <sup>[[#fn:r1371|1371]]</sup> ; Butler et al. 2009 <sup>[[#fn:r1372|1372]]</sup> ).

Drained and managed peatlands are GHG-emission hotspots (Swails et al. 2018 <sup>[[#fn:r1373|1373]]</sup> ; Hergoualc’h et al. 2017a, 2017b <sup>[[#fn:r1374|1374]]</sup> ; Roman-Cuesta et al. 2016 <sup>[[#fn:r1375|1375]]</sup> ). In most cases, lowering of the water table leads to direct and indirect CO <sub>2</sub> and N <sub>2</sub> O emissions to the atmosphere, with rates dependent on a range of factors, including the groundwater level and the water content of surface peat layers, nutrient content, temperature, and vegetation communities. The exception is nutrient-limited boreal peatlands (Minkkinen et al. 2018 <sup>[[#fn:r1376|1376]]</sup> ; Ojanen et al. 2014 <sup>[[#fn:r1377|1377]]</sup> ). Drainage also increases erosion and dissolved organic carbon loss, removing stored carbon into streams as dissolved and particulate organic carbon, which ultimately returns to the atmosphere (Moore et al. 2013 <sup>[[#fn:r1378|1378]]</sup> ; Evans et al. 2016 <sup>[[#fn:r1379|1379]]</sup> ).

In tropical peatlands, oil palm is the most widespread plantation crop and, on average, it emits around 40 tCO <sub>2</sub> ha <sup>–1</sup> yr <sup>–1</sup> ; Acacia plantations for pulpwood are the second most widespread plantation crop and emit around 73 tCO <sub>2</sub> ha <sup>–1</sup> yr <sup>–1</sup> (Drösler et al. 2013 <sup>[[#fn:r1380|1380]]</sup> ). Other land uses typically emit less than 37 tCO <sub>2</sub> ha <sup>-1</sup> yr <sup>-1</sup> . Total emissions from peatland drainage in the region are estimated to be between 0.07 and 1.1 GtCO <sub>2</sub> yr <sup>–1</sup> (Houghton and Nassikas 2017 <sup>[[#fn:r1381|1381]]</sup> ; Frolking et al. 2011 <sup>[[#fn:r1382|1382]]</sup> ). Land-use change also affects the fluxes of N <sub>2</sub> O and CH <sub>4</sub> . Undisturbed tropical peatlands emit about 0.8 MtCH <sub>4</sub> yr <sup>-1</sup> and 0.002 MtN <sub>2</sub> O yr <sup>-1</sup> , while disturbed peatlands emit 0.1 MtCH <sub>4</sub> yr <sup>–1</sup> and 0.2 MtN <sub>2</sub> O–N yr <sup>–1</sup> (Frolking et al. 2011 <sup>[[#fn:r1383|1383]]</sup> ). These N <sub>2</sub> O emissions are probably low, as new findings show that emissions from fertilised oil palm can exceed 20 kgN <sub>2</sub> O–N ha <sup>–1</sup> yr <sup>–1</sup> (Oktarita et al. 2017 <sup>[[#fn:r1384|1384]]</sup> ).

In the temperate and boreal zones, peatland drainage often leads to emissions in the order of 0.9 to 9.5 tCO <sub>2</sub> ha <sup>–1</sup> y <sup>–1</sup> in forestry plantations and 21 to 29 tCO <sub>2</sub> ha <sup>–1</sup> y <sup>–1</sup> in grasslands and croplands. Nutrient-poor sites often continue to be CO <sub>2</sub> sinks for long periods (e.g., 50 years) following drainage and, in some cases, sinks for atmospheric CH <sub>4</sub> , even when drainage ditch emissions are considered (Minkkinen et al. 2018 <sup>[[#fn:r1385|1385]]</sup> ; Ojanen et al. 2014 <sup>[[#fn:r1386|1386]]</sup> ). Undisturbed boreal and temperate peatlands emit about 30 MtCH <sub>4</sub> yr <sup>-1</sup> and 0.02 MtN <sub>2</sub> O–N yr <sup>-1</sup> , while disturbed peatlands emit 0.1 MtCH <sub>4</sub> yr <sup>–1</sup> and 0.2 MtN <sub>2</sub> O–N yr <sup>–1</sup> (Frolking et al. 2011 <sup>[[#fn:r1387|1387]]</sup> ).

Fire emissions from tropical peatlands are only a serious issue in Southeast Asia, where they are responsible for 634 (66–4070) MtCO <sub>2</sub> yr <sup>–1</sup> (van der Werf et al. 2017 <sup>[[#fn:r1388|1388]]</sup> ). Much of the variability is linked with the El Niño–Southern Oscillation (ENSO), which produces drought conditions in this region. Anomalously active fire seasons have also been observed in non-drought years and this has been attributed to the increasing effect of high temperatures that dry vegetation out during short dry spells in otherwise normal rainfall years (Fernandes et al. 2017 <sup>[[#fn:r1389|1389]]</sup> ; Gaveau et al. 2014 <sup>[[#fn:r1390|1390]]</sup> ). Fires have significant societal impacts; for example, the 2015 fires caused more than 100,000 additional deaths across Indonesia, Malaysia and Singapore, and this event was more than twice as deadly as the 2006 El Niño event (Koplitz et al. 2016 <sup>[[#fn:r1391|1391]]</sup> ).

Peatland degradation in other parts of the world differs from Asia. In Africa, for large peat deposits like those found in the Cuvette Centrale in the Congo Basin or in the Okavango inland delta, the principle threat is changing rainfall regimes due to climate variability and change (Weinzierl et al. 2016 <sup>[[#fn:r1392|1392]]</sup> ; Dargie et al. 2017 <sup>[[#fn:r1393|1393]]</sup> ). Expansion of agriculture is not yet a major factor in these regions. In the Western Amazon, extraction of non-timber forest products like the fruits of Mauritia flexuosa (moriche palm) and Suri worms are major sources of degradation that lead to losses of carbon stocks (Hergoualc’h et al. 2017a <sup>[[#fn:r1394|1394]]</sup> ).

The effects of peatland degradation on livelihoods have not been systematically characterised. In places where plantation crops are driving the conversion of peat swamps, the financial benefits can be considerable. One study in Indonesia found that the net present value of an oil palm plantation is between 3,835 and 9,630 USD per ha to land owners (Butler et al. 2009 <sup>[[#fn:r1395|1395]]</sup> ). High financial returns are creating incentives for the expansion of smallholder production in peatlands. Smallholder plantations extend over 22% of the peatlands in insular Southeast Asia compared to 27% for industrial plantations (Miettinen et al. 2016 <sup>[[#fn:r1396|1396]]</sup> ). In places where income is generated from extraction of marketable products, ecosystem degradation probably has a negative effect on livelihoods. For example, the sale of fruits of ''M. flexuosa'' in some parts of the western Amazon constitutes as much as 80% of the winter income of many rural households, but information on trade values and value chains of ''M. flexuosa'' is still sparse (Sousa et al. 2018 <sup>[[#fn:r1397|1397]]</sup> ; Virapongse et al. 2017 <sup>[[#fn:r1398|1398]]</sup> ).

There is little experience with peatland restoration in the tropics. Experience from northern latitudes suggests that extensive damage and changes in hydrological conditions mean that restoration in many cases is unachievable (Andersen et al. 2017 <sup>[[#fn:r1399|1399]]</sup> ). In the case of Southeast Asia, where peatlands form as raised bogs, drainage leads to collapse of the dome, and this collapse cannot be reversed by rewetting. Nevertheless, efforts are underway to develop solutions, or at least partial solutions in Southeast Asia, for example, by the Indonesian Peatland Restoration Agency. The first step is to restore the hydrological regime in drained peatlands, but so far experiences with canal blocking and reflooding of the peat have been only partially successful (Ritzema et al. 2014 <sup>[[#fn:r1400|1400]]</sup> ). Market incentives with certification through the Roundtable on Sustainable Palm Oil have also not been particularly successful as many concessions seek certification only after significant environmental degradation has occurred (Carlson et al. 2017 <sup>[[#fn:r1401|1401]]</sup> ). Certification had no discernible effect on forest loss or fire detection in peatlands in Indonesia. To date there is no documentation of restoration methods or successes in many other parts of the tropics. However, in situations where degradation does not involve drainage, ecological restoration may be possible. In South America, for example, there is growing interest in restoration of palm swamps, and as experiences are gained it will be important to document success factors to inform successive efforts (Virapongse et al. 2017 <sup>[[#fn:r1402|1402]]</sup> ).

In higher latitudes where degraded peatlands have been drained, the most effective option to reduce losses from these large organic carbon stocks is to change hydrological conditions and increase soil moisture and surface wetness (Regina et al. 2015 <sup>[[#fn:r1403|1403]]</sup> ). Long-term GHG monitoring in boreal sites has demonstrated that rewetting and restoration noticeably reduce emissions compared to degraded drained sites and can restore the carbon sink function when vegetation is re-established (Wilson et al. 2016 <sup>[[#fn:r1404|1404]]</sup> ; IPCC 2014a <sup>[[#fn:r1405|1405]]</sup> ; Nugent et al. 2018 <sup>[[#fn:r1406|1406]]</sup> ) although, restored ecosystems may not yet be as resilient as their undisturbed counterparts (Wilson et al. 2016 <sup>[[#fn:r1407|1407]]</sup> ). Several studies have demonstrated the co-benefits of rewetting specific degraded peatlands for biodiversity, carbon sequestration, (Parry et al. 2014 <sup>[[#fn:r1408|1408]]</sup> ; Ramchunder et al. 2012 <sup>[[#fn:r1409|1409]]</sup> ; Renou-Wilson et al. 2018 <sup>[[#fn:r1410|1410]]</sup> ) and other ecosystem services, such as improvement of water storage and quality (Martin-Ortega et al. 2014 <sup>[[#fn:r1411|1411]]</sup> ) with beneficial consequences for human well-being (Bonn et al. 2016 <sup>[[#fn:r1412|1412]]</sup> ; Parry et al. 2014 <sup>[[#fn:r1413|1413]]</sup> ).

<span id="biochar"></span>
=== 4.9.5 Biochar ===

<div id="section-4-9-5-biochar-block-1"></div>

Biochar is organic matter that is carbonised by heating in an oxygen-limited environment, and used as a soil amendment. The properties of biochar vary widely, dependent on the feedstock and the conditions of production. Biochar could make a significant contribution to mitigating both land degradation and climate change, simultaneously.

<div id="section-4-9-5-1-role-of-biochar-in-climate-change-mitigation"></div>

<span id="role-of-biochar-in-climate-change-mitigation"></span>
==== 4.9.5.1 Role of biochar in climate change mitigation ====

<div id="section-4-9-5-1-role-of-biochar-in-climate-change-mitigation-block-1"></div>

Biochar is relatively resistant to decomposition compared with fresh organic matter or compost, so represents a long-term carbon store ( ''very high confidence'' ). Biochars produced at higher temperature (&gt;450°C) and from woody material have greater stability than those produced at lower temperature (300–450°C), and from manures ( ''very high confidence'' ) (Singh et al. 2012 <sup>[[#fn:r1414|1414]]</sup> ; Wang et al. 2016b <sup>[[#fn:r1415|1415]]</sup> ). Biochar stability is influenced by soil properties: biochar carbon can be further stabilised by interaction with clay minerals and native SOM ( ''medium evidence'' ) (Fang et al. 2015 <sup>[[#fn:r1416|1416]]</sup> ). Biochar stability is estimated to range from decades to thousands of years, for different biochars in different applications (Singh et al. 2015 <sup>[[#fn:r1417|1417]]</sup> ; Wang et al. 2016 <sup>[[#fn:r1418|1418]]</sup> ). Biochar stability decreases as ambient temperature increases ( ''limited evidence'' ) (Fang et al. 2017 <sup>[[#fn:r1419|1419]]</sup> ).

Biochar can enhance soil carbon stocks through ‘negative priming’, in which rhizodeposits are stabilised through sorption of labile carbon on biochar, and formation of biochar-organo-mineral complexes (Weng et al. 2015 <sup>[[#fn:r1420|1420]]</sup> , 2017 <sup>[[#fn:r1421|1421]]</sup> , 2018 <sup>[[#fn:r1422|1422]]</sup> ; Wang et al. 2016b). Conversely, some studies show increased turnover of native soil carbon (‘positive priming’) due to enhanced soil microbial activity induced by biochar. In clayey soils, positive priming is minor and short-lived compared to negative priming effects, which dominate in the medium to long term (Singh and Cowie 2014 <sup>[[#fn:r1421|1421]]</sup> ; Wang et al. 2016b <sup>[[#fn:r1422|1422]]</sup> ). Negative priming has been observed particularly in loamy grassland soil (Ventura et al. 2015 <sup>[[#fn:r1423|1423]]</sup> ) and clay-dominated soils, whereas positive priming is reported in sandy soils (Wang et al. 2016b <sup>[[#fn:r1424|1424]]</sup> ) and those with low carbon content (Ding et al. 2018 <sup>[[#fn:r1425|1425]]</sup> ).

Biochar can provide additional climate-change mitigation by decreasing nitrous oxide (N <sub>2</sub> O) emissions from soil, due in part to decreased substrate availability for denitrifying organisms, related to the molar H/C ratio of the biochar (Cayuela et al. 2015 <sup>[[#fn:r1426|1426]]</sup> ). However, this impact varies widely: meta-analyses found an average decrease in N <sub>2</sub> O emissions from soil of 30–54%, (Cayuela et al. 2015 <sup>[[#fn:r1427|1427]]</sup> ; Borchard et al. 2019 <sup>[[#fn:r1428|1428]]</sup> ; Moore 2002 <sup>[[#fn:r1429|1429]]</sup> ), although another study found no significant reduction in field conditions when weighted by the inverse of the number of observations per site (Verhoeven et al. 2017 <sup>[[#fn:r1430|1430]]</sup> ). Biochar has been observed to reduce methane emissions from flooded soils, such as rice paddies, though, as for N <sub>2</sub> O, results vary between studies and increases have also been observed (He et al. 2017 <sup>[[#fn:r1431|1431]]</sup> ; Kammann et al. 2017 <sup>[[#fn:r1432|1432]]</sup> ). Biochar has also been found to reduce methane uptake by dryland soils, though the effect is small in absolute terms (Jeffery et al. 2016 <sup>[[#fn:r1433|1433]]</sup> ).

Additional climate benefits of biochar can arise through: reduced nitrogen fertiliser requirements, due to reduced losses of nitrogen through leaching and/or volatilisation (Singh et al. 2010 <sup>[[#fn:r1434|1434]]</sup> ) and enhanced biological nitrogen fixation (Van Zwieten et al. 2015 <sup>[[#fn:r1435|1435]]</sup> ); increased yields of crop, forage, vegetable and tree species (Biederman and Harpole 2013 <sup>[[#fn:r1436|1436]]</sup> ), particularly in sandy soils and acidic tropical soils (Simon et al. 2017 <sup>[[#fn:r1437|1437]]</sup> ); avoided GHG emissions from manure that would otherwise be stockpiled, crop residues that would be burned or processing residues that would be landfilled; and reduced GHG emissions from compost when biochar is added (Agyarko-Mintah et al. 2017 <sup>[[#fn:r1438|1438]]</sup> ; Wu et al. 2017a <sup>[[#fn:r1439|1439]]</sup> ).

Climate benefits of biochar could be substantially reduced through reduction in albedo if biochar is surface-applied at high rates to light-coloured soils (Genesio et al. 2012 <sup>[[#fn:r1440|1440]]</sup> ; Bozzi et al. 2015 <sup>[[#fn:r1441|1441]]</sup> ; Woolf et al. 2010 <sup>[[#fn:r1442|1442]]</sup> ), or if black carbon dust is released (Genesio et al. 2016 <sup>[[#fn:r1443|1443]]</sup> ). Pelletising or granulating biochar, and applying below the soil surface or incorporating into the soil, minimises the release of black carbon dust and reduces the effect on albedo (Woolf et al. 2010 <sup>[[#fn:r1444|1444]]</sup> ).

Biochar is a potential ‘negative emissions’ technology: the thermochemical conversion of biomass to biochar slows mineralisation of the biomass, delivering long-term carbon storage; gases released during pyrolysis can be combusted for heat or power, displacing fossil energy sources, and could be captured and sequestered if linked with infrastructure for CCS (Smith 2016 <sup>[[#fn:r1445|1445]]</sup> ). Studies of the lifecycle climate change impacts of biochar systems generally show emissions reduction in the range 0.4 –1.2 tCO <sub>2</sub> e t <sup>–1</sup> (dry) feedstock (Cowie et al. 2015 <sup>[[#fn:r1446|1446]]</sup> ). Use of biomass for biochar can deliver greater benefits than use for bioenergy, if applied in a context where it delivers agronomic benefits and/or reduces non-CO <sub>2</sub> GHG emissions (Ji et al. 2018 <sup>[[#fn:r1447|1447]]</sup> ; Woolf et al. 2010 <sup>[[#fn:r1448|1448]]</sup> , 2018; Xuetal.2019).A global analysis of technical potential, in which biomass supply constraints were applied to protect against food insecurity, loss of habitat and land degradation, estimated technical potential abatement of 3.7–6.6 GtCO <sub>2</sub> e yr <sup>–1</sup> (including 2.6–4.6 GtCO <sub>2</sub> e yr <sup>–1</sup> carbon stabilisation), with theoretical potential to reduce total emissions over the course of a century by 240–475 GtCO <sub>2</sub> e (Woolf et al. 2010). Fuss et al. (2018) propose a range of 0.5–2 GtCO <sub>2</sub> e per year as the sustainable potential for negative emissions through biochar. Mitigation potential of biochar is reviewed in Chapter 2.

<div id="section-4-9-5-2-role-of-biochar-in-management-of-land-degradation"></div>

<span id="role-of-biochar-in-management-of-land-degradation"></span>
==== 4.9.5.2 Role of biochar in management of land degradation ====

<div id="section-4-9-5-2-role-of-biochar-in-management-of-land-degradation-block-1"></div>

Biochars generally have high porosity, high surface area and surface-active properties that lead to high absorptive and adsorptive capacity, especially after interaction in soil (Joseph et al. 2010 <sup>[[#fn:r1450|1450]]</sup> ). As a result of these properties, biochar could contribute to avoiding, reducing and reversing land degradation through the following documented benefits:

* Improved nutrient use efficiency due to reduced leaching of nitrate and ammonium (e.g., Haider et al. 2017 <sup>[[#fn:r1451|1451]]</sup> ) and increased availability of phosphorus in soils with high phosphorus fixation capacity (Liu et al. 2018c <sup>[[#fn:r1452|1452]]</sup> ), potentially reducing nitrogen and phosphorus fertiliser requirements.
* Management of heavy metals and organic pollutants: through reduced bioavailability of toxic elements (O’Connor et al. 2018 <sup>[[#fn:r1453|1453]]</sup> ; Peng et al. 2018 <sup>[[#fn:r1454|1454]]</sup> ), by reducing availability, through immobilisation due to increased pH and redox effects (Rizwan et al. 2016 <sup>[[#fn:r1455|1455]]</sup> ) and adsorption on biochar surfaces (Zhang et al. 2013 <sup>[[#fn:r1456|1456]]</sup> ) thus providing a means of remediating contaminated soils, and enabling their utilisation for food production.
* Stimulation of beneficial soil organisms, including earthworms and mycorrhizal fungi (Thies et al. 2015 <sup>[[#fn:r1457|1457]]</sup> ).
* Improved porosity and water-holding capacity (Quin et al. 2014 <sup>[[#fn:r1458|1458]]</sup> ), particularly in sandy soils (Omondi et al. 2016 <sup>[[#fn:r1459|1459]]</sup> ), enhancing microbial function during drought (Paetsch et al. 2018 <sup>[[#fn:r1460|1460]]</sup> ).
* Amelioration of soil acidification, through application of biochars with high pH and acid-neutralising capacity (Chan et al. 2008 <sup>[[#fn:r1461|1461]]</sup> ; Van Zwieten et al. 2010 <sup>[[#fn:r1462|1462]]</sup> ).

Biochar systems can deliver a range of other co-benefits, including destruction of pathogens and weed propagules, avoidance of landfill, improved handling and transport of wastes such as sewage sludge, management of biomass residues such as environmental weeds and urban greenwaste, reduction of odours and management of nutrients from intensive livestock facilities, reduction in environmental nitrogen pollution and protection of waterways. As a compost additive, biochar has been found to reduce leaching and volatilisation of nutrients, increasing nutrient retention through absorption and adsorption processes (Joseph et al. 2018 <sup>[[#fn:r1463|1463]]</sup> ).

While many studies report positive responses, some studies have found negative or zero impacts on soil properties or plant response (e.g., Kuppusamy et al. 2016 <sup>[[#fn:r1464|1464]]</sup> ). The risk that biochar may enhance polycyclic aromatic hydrocarbon (PAH) in soil or sediments has been raised (Quilliam et al. 2013 <sup>[[#fn:r1465|1465]]</sup> ; Ojeda et al. 2016 <sup>[[#fn:r1466|1466]]</sup> ), but bioavailability of PAH in biochar has been shown to be very low (Hilber et al. 2017 <sup>[[#fn:r1467|1467]]</sup> ) Pyrolysis of biomass leads to losses of volatile nutrients, especially nitrogen. While availability of nitrogen and phosphorus in biochar is lower than in fresh biomass, (Xu et al. 2016 <sup>[[#fn:r1468|1468]]</sup> ) the impact of biochar on plant uptake is determined by the interactions between biochar, soil minerals and activity of microorganisms (e.g., Vanek and Lehmann 2015 <sup>[[#fn:r1655|1655]]</sup> ; Nguyen et al. 2017 <sup>[[#fn:r1469|1469]]</sup> ). To avoid negative responses, it is important to select biochar formulations to address known soil constraints, and to apply biochar prior to planting (Nguyen et al. 2017 <sup>[[#fn:r1470|1470]]</sup> ). Nutrient enrichment improves the performance of biochar from low nutrient feedstocks (Joseph et al. 2013 <sup>[[#fn:r1471|1471]]</sup> ). While there are many reports of biochar reducing disease or pest incidence, there are also reports of nil or negative effects (Bonanomi et al. 2015 <sup>[[#fn:r1472|1472]]</sup> ). Biochar may induce systemic disease resistance (e.g., Elad et al. 2011 <sup>[[#fn:r1473|1473]]</sup> ), though Viger et al. (2015) <sup>[[#fn:r1474|1474]]</sup> reported down-regulation of plant defence genes, suggesting increased susceptibility to insect and pathogen attack. Disease suppression where biochar is applied is associated with increased microbial diversity and metabolic potential of the rhizosphere microbiome (Kolton et al. 2017 <sup>[[#fn:r1475|1475]]</sup> ). Differences in properties related to feedstock (Bonanomi et al. 2018 <sup>[[#fn:r1476|1476]]</sup> ) and differential response to biochar dose, with lower rates more effective (Frenkel et al. 2017 <sup>[[#fn:r1477|1477]]</sup> ), contribute to variable disease responses.

The constraints on biochar adoption include: the high cost and limited availability due to limited large-scale production; limited amount of unutilised biomass; and competition for land for growing biomass. While early biochar research tended to use high rates of application (10 t ha <sup>–1</sup> or more) subsequent studies have shown that biochar can be effective at lower rates, especially when combined with chemical or organic fertilisers (Joseph et al. 2013 <sup>[[#fn:r1478|1478]]</sup> ). Biochar can be produced at many scales and levels of engineering sophistication, from simple cone kilns and cookstoves to large industrial-scale units processing several tonnes of biomass per hour (Lehmann and Joseph 2015 <sup>[[#fn:r1479|1479]]</sup> ). Substantial technological development has occurred recently, though large-scale deployment is limited to date.

Governance of biochar is required to manage climate, human health and contamination risks associated with biochar production in poorly designed or operated facilities that release methane or particulates (Downie et al. 2012 <sup>[[#fn:r1480|1480]]</sup> ; Buss et al. 2015 <sup>[[#fn:r1481|1481]]</sup> ), to ensure quality control of biochar products, and to ensure that biomass is sourced sustainably and is uncontaminated. Measures could include labelling standards, sustainability certification schemes and regulation of biochar production and use. Governance mechanisms should be tailored to context, commensurate with risks of adverse outcomes.

In summary, application of biochar to soil can improve soil chemical, physical and biological attributes, enhancing productivity and resilience to climate change, while also delivering climate-change mitigation through carbon sequestration and reduction in GHG emissions ( ''medium agreement, robust evidence'' ). However, responses to biochar depend on the biochar’s properties, which are in turn dependent on feedstock and biochar production conditions, and the soil and crop to which it is applied. Negative or nil results have been recorded.Agronomic and methane-reduction benefits appear greatest in tropical regions, where acidic soils predominate and suboptimal rates of lime and fertiliser are common, while carbon stabilisation is greater in temperate regions. Biochar is most effective when applied in low volumes to the most responsive soils and when properties are matched to the specific soil constraints and plant needs. Biochar is thus a practice that has potential to address land degradation and climate change simultaneously, while also supporting sustainable development. The potential of biochar is limited by the availability of biomass for its production. Biochar production and use requires regulation and standardisation to manage risks ( ''strong agreement'' ).

<span id="management-of-land-degradation-induced-by-tropical-cyclones"></span>
=== 4.9.6 Management of land degradation induced by tropical cyclones ===

<div id="section-4-9-6-management-of-land-degradation-induced-by-tropical-cyclones-block-1"></div>

Tropical cyclones are normal disturbances that natural ecosystems have been affected by and recovered from for millennia. Climate models mostly predict decreasing frequency of tropical cyclones, but dramatically increasing intensity of the strongest storms, as well as increasing rainfall rates (Bacmeister et al. 2018 <sup>[[#fn:r1482|1482]]</sup> ; Walsh et al. 2016b <sup>[[#fn:r1483|1483]]</sup> ). Large amplitude fluctuations in the frequency and intensity complicate both the detection and attribution of tropical cyclones to climate change (Lin and Emanuel 2016b). Yet, the force of high-intensity cyclones has increased and is expected to escalate further due to global climate change ( ''medium agreement, robust evidence'' ) (Knutson et al. 2010 <sup>[[#fn:r1484|1484]]</sup> ; Bender et al. 2010 <sup>[[#fn:r1485|1485]]</sup> ; Vecchi et al. 2008 <sup>[[#fn:r1486|1486]]</sup> ; Bhatia et al. 2018 <sup>[[#fn:r1487|1487]]</sup> ; Tu et al. 2018 <sup>[[#fn:r1488|1488]]</sup> ; Sobel et al. 2016 <sup>[[#fn:r1489|1489]]</sup> ). Tropical cyclone paths are also shifting towards the poles, increasing the area subject to tropical cyclones (Sharmila and Walsh 2018 <sup>[[#fn:r1490|1490]]</sup> ; Lin and Emanuel 2016b <sup>[[#fn:r1491|1491]]</sup> ). Climate change alone will affect the hydrology of individual wetland ecosystems, mostly through changes in precipitation and temperature regimes with great global variability (Erwin 2009 <sup>[[#fn:r1492|1492]]</sup> ). Over the last seven decades, the speed at which tropical cyclones move has decreased significantly, as expected from theory, exacerbating the damage on local communities from increasing rainfall amounts and high wind speed (Kossin 2018 <sup>[[#fn:r1493|1493]]</sup> ). Tropical cyclones will accelerate changes in coastal forest structure and composition. The heterogeneity of land degradation at coasts that are affected by tropical cyclones can be further enhanced by the interaction of its components (for example, rainfall, wind speed, and direction) with topographic and biological factors (for example, species susceptibility) (Luke et al. 2016 <sup>[[#fn:r1494|1494]]</sup> ).

Small Island Developing States (SIDS) are particularly affected by land degradation induced by tropical cyclones; recent examples are Matthew (2016) in the Caribbean, and Pam (2015) and Winston (2016) in the Pacific (Klöck and Nunn 2019 <sup>[[#fn:r1495|1495]]</sup> ; Handmer and Nalau 2019 <sup>[[#fn:r1496|1496]]</sup> ). Even if the Pacific Ocean has experienced cyclones of unprecedented intensity in recent years, their geomorphological effects may not be unprecedented (Terry and Lau 2018 <sup>[[#fn:r1497|1497]]</sup> ).

Cyclone impacts on coastal areas is not restricted to SIDS, but a problem for all low-lying coastal areas (Petzold and Magnan 2019 <sup>[[#fn:r1498|1498]]</sup> ). The Sundarbans, one of the world’s largest coastal wetlands, covers about one million hectares between Bangladesh and India. Large areas of the Sundarbans mangroves have been converted into paddy fields over the past two centuries and, more recently, into shrimp farms (Ghosh et al. 2015 <sup>[[#fn:r1499|1499]]</sup> ). In 2009, cyclone Aila caused incremental stresses on the socio-economic conditions of the Sundarbans coastal communities through rendering huge areas of land unproductive for a long time (Abdullah et al. 2016 <sup>[[#fn:r1500|1500]]</sup> ). The impact of Aila was widespread throughout the Sundarbans mangroves, showing changes between the pre- and post-cyclonic period of 20–50% in the enhanced vegetation index (Dutta et al. 2015 <sup>[[#fn:r1501|1501]]</sup> ), although the magnitude of the effects of the Sundarbans mangroves derived from climate change is not yet defined (Payo et al. 2016 <sup>[[#fn:r1502|1502]]</sup> ; Loucks et al. 2010 <sup>[[#fn:r1503|1503]]</sup> ; Gopal and Chauhan 2006 <sup>[[#fn:r1504|1504]]</sup> ; Ghosh et al. 2015 <sup>[[#fn:r1505|1505]]</sup> ; Chaudhuri et al. 2015 <sup>[[#fn:r1506|1506]]</sup> ). There is ''high agreement'' that the joint effect of climate change and land degradation will be very negative for the area, strongly affecting the environmental services provided by these forests, including the extinction of large mammal species (Loucks et al. 2010 <sup>[[#fn:r1507|1507]]</sup> ). The changes in vegetation are mainly due to inundation and erosion (Payo et al. 2016 <sup>[[#fn:r1508|1508]]</sup> ).

Tropical cyclone Nargis unexpectedly hit the Ayeyarwady River delta (Myanmar) in 2008 with unprecedented and catastrophic damages to livelihoods, destruction of forests and erosion of fields (Fritz et al. 2009 <sup>[[#fn:r1509|1509]]</sup> ) as well as eroding the shoreline 148 m compared with the long-term average (1974–2015) of 0.62 m yr <sup>-1</sup> . This is an example of the disastrous effects that changing cyclone paths can have on areas previously not affected by cyclones (Fritz et al. 2010 <sup>[[#fn:r1510|1510]]</sup> ).

<div id="section-4-9-6-1-management-of-coastal-wetlands"></div>

<span id="management-of-coastal-wetlands"></span>
==== 4.9.6.1 Management of coastal wetlands ====

<div id="section-4-9-6-1-management-of-coastal-wetlands-block-1"></div>

Tropical cyclones mainly, but not exclusively, affect coastal regions, threatening maintenance of the associated ecosystems, mangroves, wetlands, seagrasses, and so on. These areas not only provide food, water and shelter for fish, birds and other wildlife, but also provide important ecosystem services such as water-quality improvement, flood abatement and carbon sequestration (Meng et al. 2017 <sup>[[#fn:r1511|1511]]</sup> ).

Despite their importance, coastal wetlands are listed amongst the most heavily damaged of natural ecosystems worldwide. Starting in the 1990s, wetland restoration and re-creation became a ‘hotspot’ in the ecological research fields (Zedler 2000 <sup>[[#fn:r1512|1512]]</sup> ). Coastal wetland restoration and preservation is an extremely cost-effective strategy for society, for example, the preservation of coastal wetlands in the USA provides storm protection services, with a cost of 23.2 billion USD yr <sup>–1</sup> (Costanza et al. 2008 <sup>[[#fn:r1513|1513]]</sup> ).

There is a ''high agreement'' with ''medium evidence'' that the success of wetland restoration depends mainly on the flow of the water through the system, the degree to which re-flooding occurs, disturbance regimes, and the control of invasive species (Burlakova et al. 2009 <sup>[[#fn:r1514|1514]]</sup> ; López-Rosas et al. 2013 <sup>[[#fn:r1515|1515]]</sup> ). The implementation of the Ecological Mangrove Rehabilitation protocol (López-Portillo et al. 2017 <sup>[[#fn:r1516|1516]]</sup> ) that includes monitoring and reporting tasks, has been proven to deliver successful rehabilitation of wetland ecosystem services.

<div id="section-4-9-6-1-management-of-coastal-wetlands-block-2"></div>

<span id="figure-4.10"></span>
<!-- START IMG -->
<!-- IMG TITLE -->
'''Figure 4.10'''

<span id="decision-tree-showing-recommended-steps-and-tasks-to-restore-a-mangrove-wetland-based-on-original-site-conditions.-modified-from-bosire-et-al.-2008."></span>
<!-- IMG CAPTION -->
'''Decision tree showing recommended steps and tasks to restore a mangrove wetland based on original site conditions. (Modified from Bosire et al. 2008.)'''
<!-- IMG FILE -->
[[File:3efd3985210292342c41b92880022bc4 Figure-4.10-1024x667.jpg]]

Decision tree showing recommended steps and tasks to restore a mangrove wetland based on original site conditions. (Modified from Bosire et al. 2008. <sup>[[#fn:r1656|1656]]</sup> )

<!-- END IMG -->
<span id="saltwater-intrusion"></span>
=== 4.9.7 Saltwater intrusion ===

<div id="section-4-9-7-saltwater-intrusion-block-1"></div>

Current environmental changes, including climate change, have caused sea levels to rise worldwide, particularly in tropical and subtropical regions (Fasullo and Nerem 2018 <sup>[[#fn:r1517|1517]]</sup> ). Combined with scarcity of water in river channels, such rises have been instrumental in the intrusion of highly saline seawater inland, posing a threat to coastal areas and an emerging challenge to land managers and policymakers. Assessing the extent of salinisation due to sea water intrusion at a global scale nevertheless remains challenging. Wicke et al. (2011) <sup>[[#fn:r1518|1518]]</sup> suggest that across the world, approximately 1.1 Gha of land is affected by salt, with 14% of this categorised as forest, wetland or some other form of protected area. Seawater intrusion is generally caused by (i) increased tidal activity, storm surges, cyclones and sea storms due to changing climate, (ii) heavy groundwater extraction or land-use changes as a result of changes in precipitation, and droughts/floods, (iii) coastal erosion as a result of destruction of mangrove forests and wetlands, (iv) construction of vast irrigation canals and drainage networks leading to low river discharge in the deltaic region; and (v) sea level rise contaminating nearby freshwater aquifers as a result of subsurface intrusion (Uddameri et al. 2014 <sup>[[#fn:r1519|1519]]</sup> ).

The Indus Delta, located in the south-eastern coast of Pakistan near Karachi in the North Arabian Sea, is one of the six largest estuaries in the world, spanning an area of 600,000 ha. The Indus delta is a clear example of seawater intrusion and land degradation due to local as well as up-country climatic and environmental conditions (Rasul et al. 2012 <sup>[[#fn:r1520|1520]]</sup> ). Salinisation and waterlogging in the up-country areas including provinces of Punjab and Sindh is, however, caused by the irrigation network and over-irrigation (Qureshi 2011 <sup>[[#fn:r1521|1521]]</sup> ).

Such degradation takes the form of high soil salinity, inundation and waterlogging, erosion and freshwater contamination. The interannual variability of precipitation with flooding conditions in some years and drought conditions in others has caused variable river flows and sediment runoff below Kotri Barrage (about 200 km upstream of the Indus delta). This has affected hydrological processes in the lower reaches of the river and the delta, contributing to the degradation (Rasul et al. 2012 <sup>[[#fn:r1657|1657]]</sup> ).

Over 480,000 ha of fertile land is now affected by sea water intrusion, wherein eight coastal subdivisions of the districts of Badin and Thatta are mostly affected (Chandio et al. 2011 <sup>[[#fn:r1658|1658]]</sup> ). A very high intrusion rate of 0.179 ± 0.0315 km yr <sup>-1</sup> , based on the analysis of satellite data, was observed in the Indus delta during the 10 years between 2004 and 2015 (Kalhoro et al. 2016 <sup>[[#fn:r1522|1522]]</sup> ). The area of agricultural crops under cultivation has been declining, with economic losses of millions of USD (IUCN 2003 <sup>[[#fn:r1523|1523]]</sup> ). Crop yields have reduced due to soil salinity, in some places failing entirely. Soil salinity varies seasonally, depending largely on the river discharge: during the wet season (August 2014), salinity (0.18 mg L <sup>–1</sup> ) reached 24 km upstream, while during the dry season (May 2013), it reached 84 km upstream (Kalhoro et al. 2016 <sup>[[#fn:r1524|1524]]</sup> ). The freshwater aquifers have also been contaminated with sea water, rendering them unfit for drinking or irrigation purposes. Lack of clean drinking water and sanitation causes widespread diseases, of which diarrhoea is most common (IUCN 2003 <sup>[[#fn:r1525|1525]]</sup> ).

Lake Urmia in northwest Iran, the second-largest saltwater lake in the world and the habitat for endemic Iranian brine shrimp, ''Artemia urmiana'' , has also been affected by salty water intrusion. During a 17- year period between 1998 and 2014, human disruption, including agriculture and years of dam building affected the natural flow of freshwater as well as salty sea water in the surrounding area of Lake Urmia. Water quality has also been adversely affected, with salinity fluctuating over time, but in recent years reaching a maximum of 340 g L <sup>–1</sup> (similar to levels in the Dead Sea). This has rendered the underground water unfit for drinking and agricultural purposes and risky to human health and livelihoods. Adverse impacts of global climate change as well as direct human impacts have caused changes in land use, overuse of underground water resources and construction of dams over rivers, which resulted in the drying-up of the lake in large part. This condition created sand, dust and salt storms in the region which affected many sectors including agriculture, water resources, rangelands, forests and health, and generally presented desertification conditions around the lake (Karbassi et al. 2010 <sup>[[#fn:r1526|1526]]</sup> ; Marjani and Jamali 2014 <sup>[[#fn:r1527|1527]]</sup> ; Shadkam et al. 2016 <sup>[[#fn:r1528|1528]]</sup> ).

Rapid irrigation expansion in the basin has, however, indirectly contributed to inflow reduction. Annual inflow to Lake Urmia has dropped by 48% in recent years. About three-fifths of this change was caused by climate change and two-fifths by water resource development and agriculture (Karbassi et al. 2010 <sup>[[#fn:r1529|1529]]</sup> ; Marjani and Jamali 2014 <sup>[[#fn:r1530|1530]]</sup> ; Shadkam et al. 2016 <sup>[[#fn:r1531|1531]]</sup> ).

In the drylands of Mexico, intensive production of irrigated wheat and cotton using groundwater (Halvorson et al. 2003 <sup>[[#fn:r1532|1532]]</sup> ) resulted in sea water intrusion into the aquifers of La Costa de Hermosillo, a coastal agricultural valley at the centre of Sonora Desert in Northwestern Mexico. Production of these crops in 1954 was on 64,000 ha of cultivated area, increasing to 132,516 ha in 1970, but decreasing to 66,044 ha in 2009 as a result of saline intrusion from the Gulf of California (Romo-Leon et al. 2014 <sup>[[#fn:r1533|1533]]</sup> ). In 2003, only 15% of the cultivated area was under production, with around 80,000 ha abandoned due to soil salinisation whereas in 2009, around 40,000 ha was abandoned (Halvorson et al. 2003 <sup>[[#fn:r1534|1534]]</sup> ; Romo-Leon et al. 2014 <sup>[[#fn:r1535|1535]]</sup> ). Salinisation of agricultural soils could be exacerbated by climate change, as Northwestern Mexico is projected to be warmer and drier under climate change scenarios (IPCC 2013a <sup>[[#fn:r1536|1536]]</sup> ).

In other countries, intrusion of seawater is exacerbated by destruction of mangrove forests. Mangroves are important coastal ecosystems that provide spawning bed for fish, timber for building, and livelihoods to dependent communities. They also act as barriers against coastal erosion, storm surges, tropical cyclones and tsunamis (Kalhoro et al. 2017 <sup>[[#fn:r1537|1537]]</sup> ) and are among the most carbon-rich stocks on Earth (Atwood et al. 2017 <sup>[[#fn:r1538|1538]]</sup> ). They nevertheless face a variety of threats: climatic (storm surges, tidal activities, high temperatures) and human (coastal developments, pollution, deforestation, conversion to aquaculture, rice culture, oil palm plantation), leading to declines in their areas. In Pakistan, using remote sensing, the mangrove forest cover in the Indus delta decreased from 260,000 ha in 1980s to 160,000 ha in 1990 (Chandio et al. 2011 <sup>[[#fn:r1539|1539]]</sup> ). Based on remotely sensed data, a sharp decline in the mangrove area was also found in the arid coastal region of Hormozgan province in southern Iran during 1972, 1987 and 1997 (Etemadi et al. 2016 <sup>[[#fn:r1540|1540]]</sup> ). Myanmar has the highest rate (about 1% yr <sup>–1</sup> ) of mangrove deforestation in the world (Atwood et al. 2017). Regarding global loss of carbon stored in the mangrove due to deforestation, four countries exhibited high levels of loss: Indonesia (3410 GgCO <sub>2</sub> yr <sup>–1</sup> ), Malaysia (1288 GgCO <sub>2</sub> yr <sup>–1</sup> ), US (206 GgCO <sub>2</sub> yr <sup>–1</sup> ) and Brazil (186 GgCO <sub>2</sub> yr <sup>–1</sup> ). Only in Bangladesh and Guinea Bissau was there no decline in the mangrove area from 2000 to 2012 (Atwood et al. 2017 <sup>[[#fn:r1541|1541]]</sup> ).

Frequency and intensity of average tropical cyclones will continue to increase (Knutson et al. 2015 <sup>[[#fn:r1543|1543]]</sup> ) and global sea level will continue to rise. The IPCC (2013) <sup>[[#fn:r1544|1544]]</sup> projected with ''medium confidence'' that the sea level in the Asia Pacific region will rise from 0.4 to 0.6 m, depending on the emission pathway, by the end of this century. Adaptation measures are urgently required to protect the world’s coastal areas from further degradation due to saline intrusion. A viable policy framework is needed to ensure that the environmental flows to deltas in order to repulse the intruding seawater.

<span id="avoiding-coastal-maladaptation"></span>
=== 4.9.8 Avoiding coastal maladaptation ===

<div id="section-4-9-8-avoiding-coastal-maladaptation-block-1"></div>

Coastal degradation – for example, beach erosion, coastal squeeze, and coastal biodiversity loss – as a result of rising sea levels is a major concern for low lying coasts and small islands ( ''high confidence'' ). The contribution of climate change to increased coastal degradation has been well documented in AR5 (Nurse et al. 2014 <sup>[[#fn:r1545|1545]]</sup> ; Wong et al. 2014 <sup>[[#fn:r1546|1546]]</sup> ) and is further discussed in Section 4.4.1.3 as well as in the IPCC Special Report on the Ocean and Cryosphere in a Changing Climate (SROCC). However, coastal degradation can also be indirectly induced by climate change as the result of adaptation measures that involve changes to the coastal environment, for example, coastal protection measures against increased flooding and erosion due to sea level rise, and storm surges transforming the natural coast to a ‘stabilised’ coastline (Cooper and Pile 2014 <sup>[[#fn:r1547|1547]]</sup> ; French 2001 <sup>[[#fn:r1548|1548]]</sup> ). Every kind of adaptation response option is context-dependent, and, in fact, sea walls play an important role for adaptation in many places. Nonetheless, there are observed cases where the construction of sea walls can be considered ‘maladaptation’ (Barnett and O’Neill 2010 <sup>[[#fn:r1549|1549]]</sup> ; Magnan et al. 2016 <sup>[[#fn:r1659|1659]]</sup> ) by leading to increased coastal degradation, such as in the case of small islands where, due to limitations of space, coastal retreat is less of an option than in continental coastal zones. There is emerging literature on the implementation of alternative coastal protection measures and mechanisms on small islands to avoid coastal degradation induced by sea walls (e.g., Mycoo and Chadwick 2012; Sovacool 2012 <sup>[[#fn:r1551|1551]]</sup> ).

In many cases, increased rates of coastal erosion due to the construction of sea walls are the result of the negligence of local coastal morphological dynamics and natural variability as well as the interplay of environmental and anthropogenic drivers of coastal change ( ''medium evidence, high agreement'' ). Sea walls in response to coastal erosion may be ill-suited for extreme wave heights under cyclone impacts and can lead to coastal degradation by keeping overflowing sea water from flowing back into the sea, and therefore affect the coastal vegetation through saltwater intrusion, as observed in Tuvalu (Government of Tuvalu 2006 <sup>[[#fn:r1552|1552]]</sup> ; Wairiu 2017 <sup>[[#fn:r1553|1553]]</sup> ). Similarly, in Kiribati, poor construction of sea walls has resulted in increased erosion and inundation of reclaimed land (Donner 2012 <sup>[[#fn:r1554|1554]]</sup> ; Donner and Webber 2014 <sup>[[#fn:r1555|1555]]</sup> ). In the Comoros and Tuvalu, sea walls have been constructed from climate change adaptation funds and ‘often by international development organisations seeking to leave tangible evidence of their investments’ (Marino and Lazrus 2015 <sup>[[#fn:r1556|1556]]</sup> , p. 344). In these cases, they have even increased coastal erosion, due to poor planning and the negligence of other causes of coastal degradation, such as sand mining (Marino and Lazrus 2015; Betzold and Mohamed 2017 <sup>[[#fn:r1557|1557]]</sup> ; Ratter et al. 2016 <sup>[[#fn:r1558|1558]]</sup> ). On the Bahamas, the installation of sea walls as a response to coastal erosion in areas with high wave action has led to the contrary effect and has even increased sand loss in those areas (Sealey 2006 <sup>[[#fn:r1559|1559]]</sup> ). The reduction of natural buffer zones – such as beaches and dunes – due to vertical structures, such as sea walls, increased the impacts of tropical cyclones on Reunion Island (Duvat et al. 2016 <sup>[[#fn:r1560|1560]]</sup> ). Such a process of ‘coastal squeeze’ (Pontee 2013 <sup>[[#fn:r1561|1561]]</sup> ) also results in the reduction of intertidal habitat zones, such as wetlands and marshes (Zhu et al. 2010 <sup>[[#fn:r1562|1562]]</sup> ). Coastal degradation resulting from the construction of sea walls, however, is not only observed in SIDS, as described above, but also on islands in the Global North, for example, the North Atlantic (Muir et al. 2014 <sup>[[#fn:r1563|1563]]</sup> ; Young et al. 2014 <sup>[[#fn:r1564|1564]]</sup> ; Cooper and Pile 2014 <sup>[[#fn:r1565|1565]]</sup> ; Bush 2004 <sup>[[#fn:r1566|1566]]</sup> ).

The adverse effects of coastal protection measures may be avoided by the consideration of local social-ecological dynamics, including critical study of the diverse drivers of ongoing shoreline changes, and the appropriate implementation of locally adequate coastal protection options (French 200 <sup>[[#fn:r1567|1567]]</sup> 1; Duvat 2013 <sup>[[#fn:r1568|1568]]</sup> ). Critical elements for avoiding maladaptation include profound knowledge of local tidal regimes, availability of relative sea level rise scenarios and projections for extreme water levels. Moreover, the downdrift effects of sea walls need to be considered, since undefended coasts may be exposed to increased erosion (Zhu et al. 2010 <sup>[[#fn:r1569|1569]]</sup> ). In some cases, it may be possible to keep intact and restore natural buffer zones as an alternative to the construction of hard engineering solutions. Otherwise, changes in land use, building codes, or even coastal realignment can be an option in order to protect and avoid the loss of the buffer function of beaches (Duvat et al. 2016 <sup>[[#fn:r1570|1570]]</sup> ; Cooper and Pile 2014 <sup>[[#fn:r1571|1571]]</sup> ). Examples in Barbados show that combinations of hard and soft coastal protection approaches can be sustainable and reduce the risk of coastal ecosystem degradation while keeping the desired level of protection for coastal users (Mycoo and Chadwick 2012 <sup>[[#fn:r1572|1572]]</sup> ). Nature-based solutions and approaches such as ‘building with nature’ (Slobbe et al. 2013 <sup>[[#fn:r1573|1573]]</sup> ) may allow for more sustainable coastal protection mechanisms and avoid coastal degradation. Examples from the Maldives, several Pacific islands and the North Atlantic show the importance of the involvement of local communities in coastal adaptation projects, considering local skills, capacities, as well as demographic and socio-political dynamics, in order to ensure the proper monitoring and maintenance of coastal adaptation measures (Sovacool 2012 <sup>[[#fn:r1574|1574]]</sup> ; Muir et al. 2014 <sup>[[#fn:r1575|1575]]</sup> ; Young et al. 2014 <sup>[[#fn:r1576|1576]]</sup> ; Buggy and McNamara 2016 <sup>[[#fn:r1577|1577]]</sup> ; Petzold 2016 <sup>[[#fn:r1578|1578]]</sup> ).

<span id="knowledge-gaps-and-key-uncertainties"></span>