Editing IPCC:AR6/SR15/Chapter-4 (section)

=== 4.3.2 Land and Ecosystem Transitions ===

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This section assesses the feasibility of mitigation and adaptation options related to land use and ecosystems. Land transitions are grouped around agriculture and food, ecosystems and forests, and coastal systems.

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<span id="agriculture-and-food"></span>
==== 4.3.2.1 Agriculture and food ====

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In a 1.5°C world, local yields are projected to decrease in tropical regions that are major food producing areas of the world (West Africa, Southeast Asia, South Asia, and Central and northern South America) (Schleussner et al., 2016) <sup>[[#fn:r181|181]]</sup> . Some high-latitude regions may benefit from the combined effects of elevated CO <sub>2</sub> and temperature because their average temperatures are below optimal temperature for crops. In both cases there are consequences for food production and quality (Cross-Chapter Box 6 in Chapter 3 on Food Security), conservation agriculture, irrigation, food wastage, bioenergy and the use of novel technologies.

'''Food production and quality''' . Increased temperatures, including 1.5°C warming, would affect the production of cereals such as wheat and rice, impacting food security (Schleussner et al., 2016) <sup>[[#fn:r182|182]]</sup> . There is ''medium agreement'' that elevated CO <sub>2</sub> concentrations can change food composition, with implications for nutritional security (Taub et al., 2008; Högy et al., 2009; DaMatta et al., 2010; Loladze, 2014; De Souza et al., 2015) <sup>[[#fn:r183|183]]</sup> , with the effects being different depending on the region (Medek et al., 2017) <sup>[[#fn:r184|184]]</sup> .

Meta-analyses of the effects of drought, elevated CO <sub>2</sub> , and temperature conclude that at 2°C local warming and above, aggregate production of wheat, maize, and rice are expected to decrease in both temperate and tropical areas (Challinor et al., 2014) <sup>[[#fn:r185|185]]</sup> . These production losses could be lowered if adaptation measures are taken (Challinor et al., 2014) <sup>[[#fn:r186|186]]</sup> , such as developing varieties better adapted to changing climate conditions.

Adaptation options can help ensure access to sufficient, quality food. Such options include conservation agriculture, improved livestock management, increasing irrigation efficiency, agroforestry and management of food loss and waste. Complementary adaptation and mitigation options, for example, the use of climate services (Section 4.3.5), bioenergy (Section 4.3.1) and biotechnology (Section 4.4.4) can also serve to reduce emissions intensity and the carbon footprint of food production.

'''Conservation agriculture (CA)''' is a soil management approach that reduces the disruption of soil structure and biotic processes by minimising tillage. A recent meta-analysis showed that no-till practices work well in water-limited agroecosystems when implemented jointly with residue retention and crop rotation, but when used independently, may decrease yields in other situations (Pittelkow et al., 2014) <sup>[[#fn:r187|187]]</sup> . Additional climate adaptations include adjusting planting times and crop varietal selection and improving irrigation efficiency. Adaptations such as these may increase wheat and maize yields by 7–12% under climate change (Challinor et al., 2014) <sup>[[#fn:r188|188]]</sup> . CA can also help build adaptive capacity ( ''medium evidence, medium agreement'' ) (H. Smith et al., 2017; Pradhan et al., 2018) <sup>[[#fn:r189|189]]</sup> and have mitigation co-benefits through improved fertiliser use or efficient use of machinery and fossil fuels (Harvey et al., 2014; Cui et al., 2018; Pradhan et al., 2018) <sup>[[#fn:r190|190]]</sup> . CA practices can also raise soil carbon and therefore remove CO <sub>2</sub> from the atmosphere (Aguilera et al., 2013; Poeplau and Don, 2015; Vicente-Vicente et al., 2016) <sup>[[#fn:r191|191]]</sup> . However, CA adoption can be constrained by inadequate institutional arrangements and funding mechanisms (Harvey et al., 2014; Baudron et al., 2015; Li et al., 2016; Dougill et al., 2017; H. Smith et al., 2017) <sup>[[#fn:r192|192]]</sup> .

'''Sustainable intensification''' of agriculture consists of agricultural systems with increased production per unit area but with management of the range of potentially adverse impacts on the environment (Pretty and Bharucha, 2014) <sup>[[#fn:r193|193]]</sup> . Sustainable intensification can increase the efficiency of inputs and enhance health and food security (Ramankutty et al., 2018) <sup>[[#fn:r194|194]]</sup> .

'''Livestock management.''' Livestock are responsible for more GHG emissions than all other food sources. Emissions are caused by feed production, enteric fermentation, animal waste, land-use change and livestock transport and processing. Some estimates indicate that livestock supply chains could account for 7.1 GtCO <sub>2</sub> per year, equivalent to 14.5% of global anthropogenic greenhouse gas emissions (Gerber et al., 2013) <sup>[[#fn:r195|195]]</sup> . Cattle (beef, milk) are responsible for about two-thirds of that total, largely due to methane emissions resulting from rumen fermentation (Gerber et al., 2013; Opio et al., 2013) <sup>[[#fn:r196|196]]</sup> .

Despite ongoing gains in livestock productivity and volumes, the increase of animal products in global diets is restricting overall agricultural efficiency gains because of inefficiencies in the conversion of agricultural primary production (e.g., crops) in the feed-animal products pathway (Alexander et al., 2017) <sup>[[#fn:r197|197]]</sup> , offsetting the benefits of improvements in livestock production systems (Clark and Tilman, 2017) <sup>[[#fn:r198|198]]</sup> .

There is increasing agreement that overall emissions from food systems could be reduced by targeting the demand for meat and other livestock products, particularly where consumption is higher than suggested by human health guidelines. Adjusting diets to meet nutritional targets could bring large co-benefits, through GHG mitigation and improvements in the overall efficiency of food systems (Erb et al., 2009; Tukker et al., 2011; Tilman and Clark, 2014; van Dooren et al., 2014; Ranganathan et al., 2016) <sup>[[#fn:r199|199]]</sup> . Dietary shifts could contribute one-fifth of the mitigation needed to hold warming below 2°C, with one-quarter of low-cost options (Griscom et al., 2017) <sup>[[#fn:r200|200]]</sup> . There, however, remains limited evidence of effective policy interventions to achieve such large-scale shifts in dietary choices, and prevailing trends are for increasing rather than decreasing demand for livestock products at the global scale (Alexandratos and Bruinsma, 2012; OECD/FAO, 2017) <sup>[[#fn:r201|201]]</sup> . How the role of dietary shift could change in 1.5°C-consistent pathways is also not clear (see Chapter 2).

Adaptation of livestock systems can include a suite of strategies such as using different breeds and their wild relatives to develop a genetic pool resilient to climatic shocks and longer-term temperature shifts (Thornton and Herrero, 2014) <sup>[[#fn:r202|202]]</sup> , improving fodder and feed management (Bell et al., 2014; Havet et al., 2014) <sup>[[#fn:r203|203]]</sup> and disease prevention and control (Skuce et al., 2013; Nguyen et al., 2016) <sup>[[#fn:r204|204]]</sup> . Most interventions that improve the productivity of livestock systems and enhance adaptation to climate changes would also reduce the emissions intensity of food production, with significant co-benefits for rural livelihoods and the security of food supplies (Gerber et al., 2013; FAO and NZAGRC, 2017a, b, c) <sup>[[#fn:r205|205]]</sup> . Whether such reductions in emission intensity result in lower or higher absolute GHG emissions depends on overall demand for livestock products, indicating the relevance of integrating supply-side with demand-side measures within food security objectives (Gerber et al., 2013; Bajželj et al., 2014) <sup>[[#fn:r206|206]]</sup> . Transitions in livestock production systems (e.g., from extensive to intensive) can also result in significant emission reductions as part of broader land-based mitigation strategies (Havlik et al., 2014) <sup>[[#fn:r207|207]]</sup> .

Overall, there is ''high agreement'' that farm strategies that integrate mixed crop–livestock systems can improve farm productivity and have positive sustainability outcomes (Havet et al., 2014; Thornton and Herrero, 2014; Herrero et al., 2015; Weindl et al., 2015) <sup>[[#fn:r208|208]]</sup> . Shifting towards mixed crop–livestock systems is estimated to reduce agricultural adaptation costs to 0.3% of total production costs while abating deforestation by 76 Mha globally, making it a highly cost-effective adaptation option with mitigation co-benefits (Weindl et al., 2015) <sup>[[#fn:r209|209]]</sup> . Evidence from various regions supports this (Thornton and Herrero, 2015) <sup>[[#fn:r210|210]]</sup> , although the feasible scale varies between regions and systems, as well as being moderated by overall demand in specific food products. In Australia, some farmers have successfully shifted to crop–livestock systems where, each year, they allocate land and forage resources in response to climate and price trends (Bell et al., 2014) <sup>[[#fn:r211|211]]</sup> . However, there can be some unintended negative impacts of such integration, including increased burdens on women, higher requirements of capital, competing uses of crop residues (e.g., feed vs. mulching vs. carbon sequestration) and higher requirements of management skills, which can be a challenge across several low income countries (Thornton and Herrero, 2015; Thornton et al., 2018) <sup>[[#fn:r212|212]]</sup> . Finally, the feasibility of improving livestock efficiency is dependent on socio-cultural context and acceptability: there remain significant issues around widespread adoption of crossbred animals, especially by smallholders (Thornton et al., 2018) <sup>[[#fn:r213|213]]</sup> .

'''Irrigation efficiency.''' Irrigation efficiency is especially critical since water endowments are expected to change, with  20–60 Mha of global cropland being projected to revert from irrigated to rain-fed land, while other areas will receive higher precipitation in shorter time spans, thus affecting irrigation demand (Elliott et al., 2014) <sup>[[#fn:r214|214]]</sup> . While increasing irrigation system efficiency is necessary, there is mixed evidence on how to enact efficiency improvements (Fader et al., 2016; Herwehe and Scott, 2018) <sup>[[#fn:r215|215]]</sup> . Physical and technical strategies include building large-scale reservoirs or dams, renovating or deepening irrigation channels, building on-farm rainwater harvesting structures, lining ponds, channels and tanks to reduce losses through percolation and evaporation, and investing in small infrastructure such as sprinkler or drip irrigation sets (Varela-Ortega et al., 2016; Sikka et al., 2018) <sup>[[#fn:r216|216]]</sup> . Each strategy has differing costs and benefits relating to unique biophysical, social, and economic contexts. Also, increasing irrigation efficiency may foster higher dependency on irrigation, resulting in a heightened sensitivity to climate that may be maladaptive in the long term (Lindoso et al., 2014) <sup>[[#fn:r217|217]]</sup> .

Improvements in irrigation efficiency would need to be supplemented with ancillary activities, such as shifting to crops that require less water and improving soil and moisture conservation (Fader et al., 2016; Hong and Yabe, 2017; Sikka et al., 2018) <sup>[[#fn:r218|218]]</sup> . Currently, the feasibility of improving irrigation efficiency is constrained by issues of replicability across scale and sustainability over time (Burney and Naylor, 2012) <sup>[[#fn:r219|219]]</sup> , institutional barriers and inadequate market linkages (Pittock et al., 2017) <sup>[[#fn:r220|220]]</sup> .

Growing evidence suggests that investing in behavioural shifts towards using irrigation technology such as micro-sprinklers or drip irrigation, is an effective and quick adaptation strategy (Varela-Ortega et al., 2016; Herwehe and Scott, 2018; Sikka et al., 2018) <sup>[[#fn:r221|221]]</sup> as opposed to large dams which have high financial, ecological and social costs (Varela-Ortega et al., 2016) <sup>[[#fn:r222|222]]</sup> . While improving irrigation efficiency is technically feasible (R. Fishman et al., 2015) <sup>[[#fn:r223|223]]</sup> and has clear benefits for environmental values (Pfeiffer and Lin, 2014; R. Fishman et al., 2015) <sup>[[#fn:r224|224]]</sup> , feasibility is regionally differentiated as shown by examples as diverse as Kansas (Jägermeyr et al., 2015) <sup>[[#fn:r225|225]]</sup> , India (R. Fishman et al., 2015) <sup>[[#fn:r226|226]]</sup> and Africa (Pittock et al., 2017) <sup>[[#fn:r227|227]]</sup> .

'''Agroforestry.''' The integration of trees and shrubs into crop and livestock systems, when properly managed, can potentially restrict soil erosion, facilitate water infiltration, improve soil physical properties and buffer against extreme events (Lasco et al., 2014; Mbow et al., 2014; Quandt et al., 2017; Sida et al., 2018) <sup>[[#fn:r228|228]]</sup> . There is ''medium evidence'' and ''high agreement'' on the feasibility of agroforestry practices that enhance productivity, livelihoods and carbon storage (Lusiana et al., 2012; Murthy, 2013; Coulibaly et al., 2017; Sida et al., 2018) <sup>[[#fn:r229|229]]</sup> , including from indigenous production systems (Coq-Huelva et al., 2017) <sup>[[#fn:r230|230]]</sup> , with variation by region, agroforestry type, and climatic conditions (Place et al., 2012; Coe et al., 2014; Mbow et al., 2014; Iiyama et al., 2017; Abdulai et al., 2018) <sup>[[#fn:r231|231]]</sup> . Long-term studies examining the success of agroforestry, however, are rare (Coe et al., 2014; Meijer et al., 2015; Brockington et al., 2016; Zomer et al., 2016) <sup>[[#fn:r232|232]]</sup> .

The extent to which agroforestry practices employed at the farm level could be scaled up globally while satisfying growing food demand is relatively unknown. Agroforestry adoption has been relatively low and uneven (Jacobi et al., 2017; Hernández-Morcillo et al., 2018) <sup>[[#fn:r233|233]]</sup> , with constraints including the expense of establishment and lack of reliable financial support, insecure land tenure, landowner’s lack of experience with trees, complexity of management practices, fluctuating market demand and prices for different food and fibre products, the time and knowledge required for management, low intermediate benefits to offset revenue lags, and inadequate market access (Pattanayak et al., 2003; Mercer, 2004; Sendzimir et al., 2011; Valdivia et al., 2012; Coe et al., 2014; Meijer et al., 2015; Coulibaly et al., 2017; Jacobi et al., 2017) <sup>[[#fn:r234|234]]</sup> .

'''Managing food loss and waste''' . The way food is produced, processed and transported strongly influences GHG emissions. Around one-third of the food produced on the planet is not consumed (FAO, 2013) <sup>[[#fn:r235|235]]</sup> , affecting food security and livelihoods (See Cross-Chapter Box 6 on Food Security in Chapter 3). Food wastage is a combination of food loss – the decrease in mass and nutritional value of food due to poor infrastructure, logistics, and lack of storage technologies and management – and food waste that derives from inappropriate human consumption that leads to food spoilage associated with inferior quality or overproduction. Food wastage could lead to an increase in emissions estimated to 1.9–2.5 GtCO <sub>2</sub> -eq yr <sup>−1</sup> (Hiç et al., 2016) <sup>[[#fn:r236|236]]</sup> .

Decreasing food wastage has high mitigation and adaptation potential and could play an important role in land transitions towards 1.5°C, provided that reduced food waste results in lower production-side emissions rather than increased consumption (Foley et al., 2011) <sup>[[#fn:r237|237]]</sup> . There is ''medium agreement'' that a combination of individual–institutional behaviour (Refsgaard and Magnussen, 2009; Thornton and Herrero, 2014) <sup>[[#fn:r238|238]]</sup> , and improved technologies and management (Lin et al., 2013; Papargyropoulou et al., 2014) <sup>[[#fn:r239|239]]</sup> can transform food waste into products with marketable value. Institutional behaviour depends on investment and policies, which if adequately addressed could enable mitigation and adaptation co-benefits in a relatively short time.

'''Novel technologies.''' New molecular biology tools have been developed that can lead to fast and precise genome modification (De Souza et al., 2016; Scheben et al., 2016) <sup>[[#fn:r240|240]]</sup> (e.g., CRISPR Cas9; Ran et al., 2013; Schaeffer and Nakata, 2015) <sup>[[#fn:r241|241]]</sup> . Such genome editing tools may moderately assist in mitigation and adaptation of agriculture in relation to climate changes, elevated CO <sub>2</sub> , drought and flooding (DaMatta et al., 2010; De Souza et al., 2015, 2016) <sup>[[#fn:r242|242]]</sup> . These tools could contribute to developing new plant varieties that can adapt to warming of 1.5°C and overshoot, potentially avoiding some of the costs of crop shifting (Schlenker and Roberts, 2009; De Souza et al., 2016) <sup>[[#fn:r243|243]]</sup> . However, biosafety concerns and government regulatory systems can be a major barrier to the use of these tools as this increases the time and cost of turning scientific discoveries into ready applicable technologies (Andow and Zwahlen, 2006; Maghari and Ardekani, 2011) <sup>[[#fn:r244|244]]</sup> .

The strategy of reducing enteric methane emissions by ruminants through the development of inhibitors or vaccines has already been attempted with some successes, although the potential for application at scale and in different situations remains uncertain. A methane inhibitor has been demonstrated to reduce methane from feedlot systems by 30% over a 12-week period (Hristov et al., 2015) <sup>[[#fn:r245|245]]</sup> with some productivity benefits, but the ability to apply it in grazing systems will depend on further technological developments as well as costs and incentives. A vaccine could potentially modify the microbiota of the rumen and be applicable even in extensive grazing systems by reducing the presence of methanogenic micro-organisms (Wedlock et al., 2013) <sup>[[#fn:r246|246]]</sup> but has not yet been successfully demonstrated to reduce emissions in live animals. Selective breeding for lower-emitting ruminants is becoming rapidly feasible, offering small but cumulative emissions reductions without requiring substantial changes in farm systems (Pickering et al., 2015) <sup>[[#fn:r247|247]]</sup> .

Technological innovation in culturing marine and freshwater micro and macro flora has significant potential to expand food, fuel and fibre resources, and could reduce impacts on land and conventional agriculture (Greene et al., 2017) <sup>[[#fn:r248|248]]</sup> .

Technological innovation could assist in increased agricultural efficiency (e.g., via precision agriculture), decrease food wastage and genetics that enhance plant adaptation traits (Section 4.4.4). Technological and associated management improvements may be ways to increase the efficiency of contemporary agriculture to help produce enough food to cope with population increases in a 1.5°C warmer world, and help reduce the pressure on natural ecosystems and biodiversity.

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==== 4.3.2.2 Forests and other ecosystems ====

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'''Ecosystem restoration.''' Biomass stocks in tropical, subtropical, temperate and boreal biomes currently hold 1085, 194, 176, 190 Gt CO <sub>2</sub> , respectively. Conservation and restoration can enhance these natural carbon sinks (Erb et al., 2017) <sup>[[#fn:r249|249]]</sup> .

Recent studies explore options for conservation, restoration and improved land management estimating up to 23 GtCO <sub>2</sub> (Griscom et al., 2017) <sup>[[#fn:r250|250]]</sup> . Mitigation potentials are dominated by reduced rates of deforestation, reforestation and forest management, and concentrated in tropical regions (Houghton, 2013; Canadell and Schulze, 2014; Grace et al., 2014; Houghton et al., 2015; Griscom et al., 2017) <sup>[[#fn:r251|251]]</sup> . Much of the literature focuses on REDD+ (reducing emissions from deforestation and forest degradation) as an institutional mechanism. However, restoration and management activities need not be limited to REDD+, and locally adapted implementation may keep costs low, capitalize on co-benefits and ensure consideration of competing for socio-economic goals (Jantke et al., 2016; Ellison et al., 2017; Perugini et al., 2017; Spencer et al., 2017) <sup>[[#fn:r252|252]]</sup> .

Half of the estimated potential can be achieved at &lt;100 USD/tCO <sub>2</sub> ; and a third of the cost-effective potential at &lt;10 USD/tCO <sub>2</sub> (Griscom et al., 2017) <sup>[[#fn:r253|253]]</sup> . Variation of costs in projects aiming to reduce emissions from deforestation is high when considering opportunity and transaction costs (Dang Phan et al., 2014; Overmars et al., 2014; Ickowitz et al., 2017; Rakatama et al., 2017) <sup>[[#fn:r254|254]]</sup> .

However, the focus on forests raises concerns of cross-biome leakage ( ''medium evidence, low agreement'' ) (Popp et al., 2014a; Strassburg et al., 2014; Jayachandran et al., 2017) <sup>[[#fn:r255|255]]</sup> and encroachment on other ecosystems (Veldman et al., 2015) <sup>[[#fn:r256|256]]</sup> . Reducing rates of deforestation constrains the land available for agriculture and grazing, with trade-offs between diets, higher yields and food prices (Erb et al., 2016a; Kreidenweis et al., 2016) <sup>[[#fn:r257|257]]</sup> . Forest restoration and conservation are compatible with biodiversity (Rey Benayas et al., 2009; Jantke et al., 2016) <sup>[[#fn:r258|258]]</sup> and available water resources; in the tropics, reducing rates of deforestation maintains cooler surface temperatures (Perugini et al., 2017) <sup>[[#fn:r259|259]]</sup> and rainfall (Ellison et al., 2017) <sup>[[#fn:r260|260]]</sup> .

Its multiple potential co-benefits have made REDD+ important for local communities, biodiversity and sustainable landscapes (Ngendakumana et al., 2017; Turnhout et al., 2017) <sup>[[#fn:r261|261]]</sup> . There is ''low agreement'' on whether climate impacts will reverse mitigation benefits of restoration (Le Page et al., 2013) <sup>[[#fn:r262|262]]</sup> by increasing the likelihood of disturbance (Anderegg et al., 2015) <sup>[[#fn:r263|263]]</sup> , or reinforce them through carbon fertilization (P. Smith et al., 2014) <sup>[[#fn:r264|264]]</sup> .

Emerging regional assessments offer new perspectives for upscaling. Strengthening coordination, additional funding sources, and access and disbursement points increase the potential of REDD+ in working towards 2°C and 1.5°C limits (Well and Carrapatoso, 2017) <sup>[[#fn:r265|265]]</sup> . While there are indications that land tenure has a positive impact (Sunderlin et al., 2014) <sup>[[#fn:r266|266]]</sup> , a meta-analysis by Wehkamp et al. (2018a) <sup>[[#fn:r267|267]]</sup> shows that there is ''medium evidence'' and ''low agreement'' on which aspects of governance improvements are supportive of conservation. Local benefits, especially for indigenous communities, will only be accrued if land tenure is respected and legally protected, which is not often the case (Sunderlin et al., 2014; Brugnach et al., 2017) <sup>[[#fn:r268|268]]</sup> . Although payments for reduced rates of deforestation may benefit the poor, the most vulnerable populations could have limited, uneven access (Atela et al., 2014) <sup>[[#fn:r269|269]]</sup> and face lower opportunity costs from deforestation (Ickowitz et al., 2017) <sup>[[#fn:r270|270]]</sup> .

'''Community-based adaptation (CbA).''' There is ''medium evidence'' and ''high agreement'' for the use of CbA. The specific actions to take will depend upon the location, context, and vulnerability of the specific community. CbA is defined as ‘a community-led process, based on communities’ priorities, needs, knowledge, and capacities, which aim to empower people to plan for and cope with the impacts of climate change’ (Reid et al., 2009) <sup>[[#fn:r271|271]]</sup> . The integration of CbA with ecosystems-based adaptation (EbA) has been increasingly promoted, especially in efforts to alleviate poverty (Mannke, 2011; Reid, 2016) <sup>[[#fn:r272|272]]</sup> .

Despite the potential and advantages of both CbA and EbA, including knowledge exchange, information access and increased social capital and equity; institutional and governance barriers still constitute a challenge for local adaptation efforts (Wright et al., 2014; Fernández-Giménez et al., 2015) <sup>[[#fn:r273|273]]</sup> .

'''Wetland management.''' In wetland ecosystems, temperature rise has direct and irreversible impacts on species functioning and distribution, ecosystem equilibrium and services, and second-order impacts on local livelihoods (see Chapter 3, Section 3.4.3). The structure and function of wetland systems are changing due to climate change. Wetland management strategies, including adjustments in infrastructural, behavioural, and institutional practices have clear implications for adaptation (Colloff et al., 2016b; Finlayson et al., 2017; Wigand et al., 2017) <sup>[[#fn:r274|274]]</sup>

Despite international initiatives on wetland restoration and management through the Ramsar Convention on Wetlands, policies have not been effective (Finlayson, 2012; Finlayson et al., 2017) <sup>[[#fn:r275|275]]</sup> . Institutional reform, such as flexible, locally relevant governance, drawing on principles of adaptive co-management, and multi-stakeholder participation becomes increasingly necessary for effective wetland management (Capon et al., 2013; Finlayson et al., 2017) <sup>[[#fn:r276|276]]</sup> .

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==== 4.3.2.3 Coastal systems ====

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'''Managing coastal stress.''' Particularly to allow for the landward relocation of coastal ecosystems under a transition to a 1.5°C warmer world, planning for climate change would need to be integrated with the use of coastlines by humans (Saunders et al., 2014; Kelleway et al., 2017) <sup>[[#fn:r277|277]]</sup> . Adaptation options for managing coastal stress include coastal hardening through the building of seawalls and the re-establishment of coastal ecosystems such as mangroves (André et al., 2016; Cooper et al., 2016) <sup>[[#fn:r278|278]]</sup> . While the feasibility of the solutions is high, they are expensive to scale ( ''robust evidence, medium agreement'' ).

There is ''low evidence'' and ''high agreement'' that reducing the impact of local stresses (Halpern et al., 2015) <sup>[[#fn:r279|279]]</sup> will improve the resilience of marine ecosystems as they transition to a 1.5°C world (O’Leary et al., 2017) <sup>[[#fn:r280|280]]</sup> .  Approaches to reducing local stresses are considered feasible, cost-effective and highly scalable. Ecosystem resilience may be increased through alternative livelihoods (e.g., sustainable aquaculture), which are among a suite of options for building resilience in coastal ecosystems. These options enjoy high levels of feasibility yet are expensive, which stands in the way of scalability ( ''robust evidence, medium agreement'' ) (Hiwasaki et al., 2015; Brugnach et al., 2017) <sup>[[#fn:r281|281]]</sup> .

Working with coastal communities has the potential for improving the resilience of coastal ecosystems. Combined with the advantages of using indigenous knowledge to guide transitions, solutions can be more effective when undertaken in partnership with local communities, cultures, and knowledge (See Box 4.3).

'''Restoration of coastal ecosystems and fisheries.''' Marine restoration is expensive compared to terrestrial restoration, and the survival of projects is currently low, with success depending on the ecosystem and site, rather than the size of the financial investment (Bayraktarov et al., 2016) <sup>[[#fn:r282|282]]</sup> . Mangrove replanting shows evidence of success globally, with numerous examples of projects that have established forests (Kimball et al., 2015; Bayraktarov et al., 2016) <sup>[[#fn:r283|283]]</sup> .

Efforts with reef-building corals have been attempted with a low level of success (Bayraktarov et al., 2016) <sup>[[#fn:r284|284]]</sup> . Technologies to help re-establish coral communities are limited (Rinkevich, 2014) <sup>[[#fn:r285|285]]</sup> , as are largely untested disruptive technologies (e.g., genetic manipulation, assisted evolution) (van Oppen et al., 2015) <sup>[[#fn:r286|286]]</sup> . Current technologies also have trouble scaling given the substantial costs and investment required (Bayraktarov et al., 2016) <sup>[[#fn:r287|287]]</sup> .

Johannessen and Macdonald (2016) <sup>[[#fn:r288|288]]</sup> report the ‘blue carbon’ sink to be 0.4–0.8% of global anthropogenic emissions. However, this does not adequately account for post-depositional processes and could overestimate removal potentials, subject to a risk of reversal. Seagrass beds will thus not contribute significantly to enabling 1.5°C-consistent pathways.

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