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

=== 4.2.1 Processes of land degradation ===

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A large array of interactive physical, chemical, biological and human processes lead to what we define in this report as land degradation (Johnson and Lewis 2007 <sup>[[#fn:r138|138]]</sup> ). The biological productivity, ecological integrity (which encompasses both functional and structural attributes of ecosystems) or the human value (which includes any benefit that people get from the land) of a given territory can deteriorate as the result of processes triggered at scales that range from a single furrow (e.g., water erosion under cultivation) to the landscape level (e.g., salinisation through raising groundwater levels under irrigation). While pressures leading to land degradation are often exerted on specific components of the land systems (i.e., soils, water, biota), once degradation processes start, other components become affected through cascading and interactive effects. For example, different pressures and degradation processes can have convergent effects, as can be the case of overgrazing leading to wind erosion, landscape drainage resulting in wetland drying, and warming causing more frequent burning; all of which can independently lead to reductions of the soil organic matter (SOM) pools as a second-order process. Still, the reduction of organic matter pools is also a first-order process triggered directly by the effects of rising temperatures (Crowther et al. 2016 <sup>[[#fn:r139|139]]</sup> ) as well as other climate changes such as precipitation shifts (Viscarra Rossel et al. 2014 <sup>[[#fn:r140|140]]</sup> ). Beyond this complexity, a practical assessment of the major land degradation processes helps to reveal and categorise the multiple pathways in which climate change exerts a degradation pressure (Table 4.1).

Conversion of freshwater wetlands to agricultural land has historically been a common way of increasing the area of arable land. Despite the small areal extent – about 1% of the earth’s surface (Hu et al. 2017 <sup>[[#fn:r141|141]]</sup> ; Dixon et al. 2016 <sup>[[#fn:r142|142]]</sup> ) – freshwater wetlands provide a very large number of ecosystem services, such as groundwater replenishment, flood protection and nutrient retention, and are biodiversity hotspots (Reis et al. 2017 <sup>[[#fn:r143|143]]</sup> ; Darrah et al. 2019 <sup>[[#fn:r144|144]]</sup> ; Montanarella et al. 2018 <sup>[[#fn:r145|145]]</sup> ). The loss of wetlands since 1900 has been estimated at about 55% globally (Davidson 2014 <sup>[[#fn:r146|146]]</sup> ) ( ''low confidence'' ) and 35% since 1970 (Darrah et al. 2019 <sup>[[#fn:r147|147]]</sup> ) ( ''medium confidence'' ) which in many situations pose a problem for adaptation to climate change. Drainage causes loss of wetlands, which can be exacerbated by climate change, further reducing the capacity to adapt to climate change (Barnett et al. 2015 <sup>[[#fn:r148|148]]</sup> ; Colloff et al. 2016 <sup>[[#fn:r149|149]]</sup> ; Finlayson et al. 2017 <sup>[[#fn:r150|150]]</sup> ) ( ''high confidence'' ).

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==== 4.2.1.1 Types of land degradation processes ====

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Land degradation processes can affect the soil, water or biotic components of the land as well as the reactions between them (Table 4.1). Across land degradation processes, those affecting the soil have received more attention. The most widespread and studied land degradation processes affecting soils are water and wind erosion, which have accompanied agriculture since its onset and are still dominant (Table 4.1). Degradation through erosion processes is not restricted to soil loss in detachment areas but includes impacts on transport and deposition areas as well (less commonly, deposition areas can have their soils improved by these inputs). Larger-scale degradation processes related to the whole continuum of soil erosion, transport and deposition include dune field expansion/ displacement, development of gully networks and the accumulation of sediments in natural and artificial water-bodies (siltation) (Poesen and Hooke 1997 <sup>[[#fn:r151|151]]</sup> ; Ravi et al. 2010 <sup>[[#fn:r152|152]]</sup> ). Long-distance sediment transport during erosion events can have remote effects on land systems, as documented for the fertilisation effect of African dust on the Amazon (Yu et al. 2015 <sup>[[#fn:r153|153]]</sup> ).

Coastal erosion represents a special case among erosional processes, with reports linking it to climate change. While human interventions in coastal areas (e.g., expansion of shrimp farms) and rivers (e.g., upstream dams cutting coastal sediment supply), and economic activities causing land subsidence (Keogh and Törnqvist 2019 <sup>[[#fn:r154|154]]</sup> ; Allison et al. 2016 <sup>[[#fn:r155|155]]</sup> ) are dominant human drivers, storms and sea-level rise have already left a significant global imprint on coastal erosion (Mentaschi et al. 2018 <sup>[[#fn:r156|156]]</sup> ). Recent projections that take into account geomorphological and socioecological feedbacks suggest that coastal wetlands may not be reduced by sea level rise if their inland growth is accommodated with proper management actions (Schuerch et al. 2018 <sup>[[#fn:r157|157]]</sup> ).

Other physical degradation processes in which no material detachment and transport are involved include soil compaction, hardening, sealing and any other mechanism leading to the loss of porous space crucial for holding and exchanging air and water (Hamza and Anderson 2005 <sup>[[#fn:r158|158]]</sup> ). A very extreme case of degradation through pore volume loss, manifested at landscape or larger scales, is ground subsidence. Typically caused by the lowering of groundwater or oil levels, subsidence involves a sustained collapse of the ground

surface, which can lead to other degradation processes such as salinisation and permanent flooding. Chemical soil degradation processes include relatively simple changes, like nutrient depletion resulting from the imbalance of nutrient extraction on harvested products and fertilisation, and more complex ones, such as acidification and increasing metal toxicity. Acidification in croplands is increasingly driven by excessive nitrogen fertilisation and, to a lower extent, by the depletion of cation like calcium, potassium or magnesium through exports in harvested biomass (Guo et al. 2010 <sup>[[#fn:r159|159]]</sup> ). One of the most relevant chemical degradation processes of soils in the context of climate change is the depletion of its organic matter pool. Reduced in agricultural soils through the increase of respiration rates by tillage and the decline of below-ground plant biomass inputs, SOM pools have been diminished also by the direct effects of warming, not only in cultivated land, but also under natural vegetation (Bond-Lamberty et al. 2018 <sup>[[#fn:r160|160]]</sup> ). Debate persists, however, on whether in more humid and carbon-rich ecosystems the simultaneous stimulation of decomposition and productivity may result in the lack of effects on soil carbon (Crowther et al. 2016 <sup>[[#fn:r161|161]]</sup> ; van Gestel et al. 2018 <sup>[[#fn:r162|162]]</sup> ). In the case of forests, harvesting – particularly if it is exhaustive, as in the case of the use of residues for energy generation – can also lead to organic matter declines (Achat et al. 2015 <sup>[[#fn:r163|163]]</sup> ). Many other degradation processes (e.g., wildfire increase, salinisation) have negative effects on other pathways of soil degradation (e.g., reduced nutrient availability, metal toxicity). SOM can be considered a ‘hub’ of degradation processes and a critical link with the climate system (Minasny et al. 2017 <sup>[[#fn:r164|164]]</sup> ).

Land degradation processes can also start from alterations in the hydrological system that are particularly important in the context of climate change. Salinisation, although perceived and reported in soils, is typically triggered by water table-level rises, driving salts to the surface under dry to sub-humid climates (Schofield and Kirkby 2003 <sup>[[#fn:r165|165]]</sup> ). While salty soils occur naturally under these climates (primary salinity), human interventions have expanded their distribution, secondary salinity with irrigation without proper drainage being the predominant cause of salinisation (Rengasamy 2006 <sup>[[#fn:r166|166]]</sup> ). Yet, it has also taken place under non-irrigated conditions where vegetation changes (particularly dry forest clearing and cultivation) have reduced the magnitude and depth of soil water uptake, triggering water table rises towards the surface. Changes in evapotranspiration and rainfall regimes can exacerbate this process (Schofield and Kirkby 2003 <sup>[[#fn:r167|167]]</sup> ). Salinisation can also result from the intrusion of sea water into coastal areas, both as a result of sea level rise and ground subsidence (Colombani et al. 2016 <sup>[[#fn:r168|168]]</sup> ).

Recurring flood and waterlogging episodes (Bradshaw et al. 2007 <sup>[[#fn:r169|169]]</sup> ; Poff 2002 <sup>[[#fn:r170|170]]</sup> ), and the more chronic expansion of wetlands over dryland ecosystems, are mediated by the hydrological system, on occasions aided by geomorphological shifts as well (Kirwan et al. 2011 <sup>[[#fn:r171|171]]</sup> ). This is also the case for the drying of continental water bodies and wetlands, including the salinisation and drying of lakes and inland seas (Anderson et al. 2003 <sup>[[#fn:r172|172]]</sup> ; Micklin 2010 <sup>[[#fn:r173|173]]</sup> ; Herbert et al. 2015 <sup>[[#fn:r174|174]]</sup> ). In the context of climate change, the degradation of peatland ecosystems is particularly relevant given their very high carbon storage and their sensitivity to changes in soils, hydrology and/or vegetation (Leifeld and Menichetti 2018 <sup>[[#fn:r175|175]]</sup> ). Drainage for land-use conversion together with peat mining are major drivers of peatland degradation, yet other factors such as the extractive use of their natural vegetation and the interactive effects of water table levels and fires (both sensitive to climate change) are important (Hergoualc’h et al. 2017a <sup>[[#fn:r176|176]]</sup> ; Lilleskov et al. 2019 <sup>[[#fn:r177|177]]</sup> ).

The biotic components of the land can also be the focus of degradation processes. Vegetation clearing processes associated with land-use changes are not limited to deforestation but include other natural and seminatural ecosystems such as grasslands (the most cultivated biome on Earth), as well as dry steppes and shrublands, which give place to croplands, pastures, urbanisation or just barren land. This clearing process is associated with net carbon losses from the vegetation and soil pool. Not all biotic degradation processes involve biomass losses. Woody encroachment of open savannahs involves the expansion of woody plant cover and/or density over herbaceous areas and often limits the secondary productivity of rangelands (Asner et al. 2004 <sup>[[#fn:r178|178]]</sup> ; Anadon et al. 2014 <sup>[[#fn:r179|179]]</sup> ). These processes have accelerated since the mid-1800s over most continents (Van Auken 2009 <sup>[[#fn:r180|180]]</sup> ). Change in plant composition of natural or semi-natural ecosystems without any significant vegetation structural changes is another pathway of degradation affecting rangelands and forests. In rangelands, selective grazing and its interaction with climate variability and/or fire can push ecosystems to new compositions with lower forage value and a higher proportion of invasive species (Illius and O ́Connor 1999 <sup>[[#fn:r181|181]]</sup> ; Sasaki et al. 2007 <sup>[[#fn:r182|182]]</sup> ), in some cases with higher carbon sequestration potential, yet with very complex interactions between vegetation and soil carbon shifts (Piñeiro et al. 2010 <sup>[[#fn:r183|183]]</sup> ). In forests, extractive logging can be a pervasive cause of degradation, leading to long-term impoverishment and, in extreme cases, a full loss of the forest cover through its interaction with other agents such as fires (Foley et al. 2007 <sup>[[#fn:r184|184]]</sup> ) or progressive intensification of land use. Invasive alien species are another source of biological degradation. Their arrival into cultivated systems is constantly reshaping crop production strategies, making agriculture unviable on occasions. In natural and seminatural systems such as rangelands, invasive plant species not only threaten livestock production through diminished forage quality, poisoning and other deleterious effects, but have cascading effects on other processes such as altered fire regimes and water cycling (Brooks et al. 2004 <sup>[[#fn:r185|185]]</sup> ). In forests, invasions affect primary productivity and nutrient availability, change fire regimes, and alter species composition, resulting in long-term impacts on carbon pools and fluxes (Peltzer et al. 2010 <sup>[[#fn:r186|186]]</sup> ).

Other biotic components of ecosystems have been shown as a focus of degradation processes. Invertebrate invasions in continental waters can exacerbate other degradation processes such as eutrophication, which is the over-enrichment of nutrients, leading to excessive algal growth (Walsh et al. 2016a <sup>[[#fn:r187|187]]</sup> ). Shifts in soil microbial and mesofaunal composition – which can be caused by pollution with pesticides or nitrogen deposition and by vegetation or disturbance regime shifts – alter many soil functions, including respiration rates and carbon release to the atmosphere (Hussain et al. 2009 <sup>[[#fn:r188|188]]</sup> ; Crowther et al. 2015 <sup>[[#fn:r189|189]]</sup> ). The role of the soil biota in modulating the effects of climate change on soil carbon has been recently demonstrated (Ratcliffe et al. 2017 <sup>[[#fn:r190|190]]</sup> ), highlighting the importance of this lesser-known component of the biota as a focal point of land degradation. Of special relevance as both indicators and agents of land degradation recovery are mycorrhiza, which are root-associated fungal organisms (Asmelash et al. 2016 <sup>[[#fn:r191|191]]</sup> ; Vasconcellos et al. 2016 <sup>[[#fn:r192|192]]</sup> ). In natural dry ecosystems, biological soil crusts composed of a broad range of organisms, including mosses, are a particularly sensitive focus for degradation (Field et al. 2010 <sup>[[#fn:r193|193]]</sup> ) with evidenced sensitivity to climate change (Reed et al. 2012 <sup>[[#fn:r194|194]]</sup> ).

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==== 4.2.1.2 Land degradation processes and climate change ====

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While the subdivision of individual processes is challenged by their strong interconnectedness, it provides a useful setting to identify the most important ‘focal points’ of climate change pressures on land degradation. Among land degradation processes, those responding more directly to climate change pressures include all types of erosion and SOM declines (soil focus), salinisation, sodification and permafrost thawing (soil/water focus), waterlogging of dry ecosystems and drying of wet ecosystems (water focus), and a broad group of biologically-mediated processes like woody encroachment, biological invasions, pest outbreaks (biotic focus), together with biological soil crust destruction and increased burning (soil/biota focus) (Table 4.1). Processes like ground subsidence can be affected by climate change indirectly through sea level rise (Keogh and Törnqvist 2019 <sup>[[#fn:r195|195]]</sup> ).

Even when climate change exerts a direct pressure on degradation processes, it can be a secondary driver subordinated to other overwhelming human pressures. Important exceptions are three processes in which climate change is a dominant global or regional pressure and the main driver of their current acceleration. These are: coastal erosion as affected by sea level rise and increased storm frequency/intensity ( ''high agreement, medium evidence'' ) (Johnson et al. 2015 <sup>[[#fn:r196|196]]</sup> ; Alongi 2015 <sup>[[#fn:r197|197]]</sup> ; Harley et al. 2017 <sup>[[#fn:r198|198]]</sup> ; Nicholls et al. 2016 <sup>[[#fn:r199|199]]</sup> ); permafrost thawing responding to warming ( ''high agreement, robust evidence'' ) (Liljedahl et al. 2016 <sup>[[#fn:r200|200]]</sup> ; Peng et al. 2016 <sup>[[#fn:r201|201]]</sup> ; Batir et al. 2017 <sup>[[#fn:r202|202]]</sup> ); and increased burning responding to warming and altered precipitation regimes ( ''high agreement, robust evidence'' ) (Jolly et al. 2015 <sup>[[#fn:r203|203]]</sup> ; Abatzoglou and Williams 2016 <sup>[[#fn:r204|204]]</sup> ; Taufik et al. 2017 <sup>[[#fn:r205|205]]</sup> ; Knorr et al. 2016 <sup>[[#fn:r206|206]]</sup> ). The previous assessment highlights the fact that climate change not only exacerbates many of the well-acknowledged ongoing land degradation processes of managed ecosystems (i.e., croplands and pastures), but becomes a dominant pressure that introduces novel degradation pathways in natural and seminatural ecosystems. Climate change has influenced species invasions and the degradation that they cause by enhancing the transport, colonisation, establishment, and ecological impact of the invasive species, and also by impairing their control practices ( ''medium agreement, medium evidence'' ) (Hellmann et al. 2008 <sup>[[#fn:r207|207]]</sup> ).

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'''Table 4.1'''

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'''Major land degradation processes and their connections with climate change.''''

For each process a ‘focal point’ (soil, water, biota) on which degradation occurs in the first place is indicated, acknowledging that most processes propagate to other land components and cascade into or interact with some of the other processes listed below. The impact of climate change on each process is categorised based on the proximity (very direct = high, very indirect = low) and dominance (dominant = high, subordinate to other pressures = low) of effects. The major effects of climate change on each process are highlighted together with the predominant pressures from other drivers. Feedbacks of land degradation processes on climate change are categorised according to the intensity (very intense = high, subtle = low) of the chemical (GHG emissions or capture) or physical (energy and momentum exchange, aerosol emissions) effects. Warming effects are indicated in red and cooling effects in blue. Specific feedbacks on climate change are highlighted.
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References in Table 4.1: (1) Bärring et al. 2003 <sup>[[#fn:r1580|1580]]</sup> ; Munson et al. 2011 <sup>[[#fn:r1581|1581]]</sup> ; Sheffield et al. 2012 <sup>[[#fn:r1582|1582]]</sup> , (2) Nearing et al. 2004 <sup>[[#fn:r1583|1583]]</sup> ; Shakesby 2011 <sup>[[#fn:r1584|1584]]</sup> ; Panthou et al. 2014 <sup>[[#fn:r1585|1585]]</sup> , (3) Johnson et al. 2015 <sup>[[#fn:r1586|1586]]</sup> ; Alongi 2015 <sup>[[#fn:r1587|1587]]</sup> ; Harley et al. 2017 <sup>[[#fn:r1588|1588]]</sup> , (4) Bond-Lamberty et al. 2018 <sup>[[#fn:r1589|1589]]</sup> ; Crowther et al. 2016 <sup>[[#fn:r1590|1590]]</sup> ; van Gestel et al. 2018 <sup>[[#fn:r1591|1591]]</sup> , (5) Colombani et al. 2016 <sup>[[#fn:r1592|1592]]</sup> , (6) Schofield and Kirkby 2003 <sup>[[#fn:r1593|1593]]</sup> ; Aragüés et al. 2015 <sup>[[#fn:r1594|1594]]</sup> ; Benini et al. 2016 <sup>[[#fn:r1595|1595]]</sup> , (7) Jobbágy et al. 2017 <sup>[[#fn:r1596|1596]]</sup> , (8) Liljedahl et al. 2016 <sup>[[#fn:r1597|1597]]</sup> ; Peng et al. 2016 <sup>[[#fn:r1598|1598]]</sup> ; Batir et al. 2017 <sup>[[#fn:r1599|1599]]</sup> , (9) Piovano et al. 2004 <sup>[[#fn:r1600|1600]]</sup> ; Osland et al. 2016 <sup>[[#fn:r1601|1601]]</sup> , (10) Burkett and Kusler 2000 <sup>[[#fn:r1602|1602]]</sup> ; Nielsen and Brock 2009 <sup>[[#fn:r1603|1603]]</sup> ; Johnson et al. 2015 <sup>[[#fn:r1604|1604]]</sup> ; Green et al. 2017 <sup>[[#fn:r1605|1605]]</sup> , (11) Panthou et al. 2014 <sup>[[#fn:r1606|1606]]</sup> ; Arnell and Gosling 2016 <sup>[[#fn:r1607|1607]]</sup> ; Vitousek et al. 2017 <sup>[[#fn:r1608|1608]]</sup> , (12) Van Auken 2009 <sup>[[#fn:r1609|1609]]</sup> ; Wigley et al. 2010 <sup>[[#fn:r1610|1610]]</sup> , (13) Vincent et al. 2014 <sup>[[#fn:r1611|1611]]</sup> ; Gonzalez et al. 2010 <sup>[[#fn:r1612|1612]]</sup> ; Scheffers et al. 2016 <sup>[[#fn:r1613|1613]]</sup> , (14) Pritchard 2011 <sup>[[#fn:r1614|1614]]</sup> ; Ratcliffe et al. 2017 <sup>[[#fn:r1615|1615]]</sup> , (15) Reed et al. 2012 <sup>[[#fn:r1616|1616]]</sup> ; Maestre et al. 2013 <sup>[[#fn:r1617|1617]]</sup> , (16) Hellmann et al. 2008 <sup>[[#fn:r1618|1618]]</sup> ; Hulme 2017 <sup>[[#fn:r1619|1619]]</sup> , (17) Pureswaran et al. 2015 <sup>[[#fn:r1620|1620]]</sup> ; Cilas et al. 2016 <sup>[[#fn:r1621|1621]]</sup> ; Macfadyen et al. 2018 <sup>[[#fn:r1622|1622]]</sup> , (18) Jolly et al. 2015 <sup>[[#fn:r1623|1623]]</sup> ; Abatzoglou and Williams 2016 <sup>[[#fn:r1624|1624]]</sup> ; Taufik et al. 2017 <sup>[[#fn:r1625|1625]]</sup> ; Knorr et al. 2016 <sup>[[#fn:r1626|1626]]</sup> , (19) Davin et al. 2010 <sup>[[#fn:r1627|1627]]</sup> ; Pinty et al. 2011 <sup>[[#fn:r1628|1628]]</sup> , (20) Wang et al. 2017b <sup>[[#fn:r1629|1629]]</sup> ; Chappell et al. 2016 <sup>[[#fn:r1630|1630]]</sup> , (21) Pendleton et al. 2012 <sup>[[#fn:r1631|1631]]</sup> , (22) Oertel et al. 2016 <sup>[[#fn:r1632|1632]]</sup> , (23) Houghton et al. 2012 <sup>[[#fn:r1633|1633]]</sup> ; Eglin et al. 2010 <sup>[[#fn:r1634|1634]]</sup> , (24) Schuur et al. 2015 <sup>[[#fn:r1635|1635]]</sup> ; Christensen et al. 2004 <sup>[[#fn:r1636|1636]]</sup> ; Walter Anthony et al. 2016 <sup>[[#fn:r1637|1637]]</sup> ; Abbott et al. 2016 <sup>[[#fn:r1638|1638]]</sup> , (25) Belnap, Walker, Munson &amp; Gill, 2014 <sup>[[#fn:r1639|1639]]</sup> ; Rutherford et al. 2017 <sup>[[#fn:r1640|1640]]</sup> , (26) Page et al. 2002 <sup>[[#fn:r1641|1641]]</sup> ; Pellegrini et al. 2018 <sup>[[#fn:r1642|1642]]</sup> .

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