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

=== 4.6.1 Impact on greenhouse gases (GHGs) ===

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Land degradation processes with direct impact on soil and terrestrial biota have great relevance in terms of CO <sub>2</sub> exchange with the atmosphere, given the magnitude and activity of these reservoirs in the global carbon cycle. As the most widespread form of soil degradation, erosion detaches the surface soil material, which typically hosts the highest organic carbon stocks, favouring the mineralisation and release as CO <sub>2</sub> . Yet complementary processes such as carbon burial may compensate for this effect, making soil erosion a long-term carbon sink ( ''low agreement, limited evidence'' ), (Wang et al. (2017b) <sup>[[#fn:r783|783]]</sup> , but see also Chappell et al. (2016) <sup>[[#fn:r784|784]]</sup> ). Precise estimation of the CO <sub>2</sub> released from eroded lands is challenged by the fact that only a fraction of the detached carbon is eventually lost to the atmosphere. It is important to acknowledge that a substantial fraction of the eroded material may preserve its organic carbon load in field conditions. Moreover, carbon sequestration may be favoured through the burial of both the deposited material and the surface of its hosting soil at the deposition location (Quinton et al. 2010 <sup>[[#fn:r785|785]]</sup> ). The cascading effects of erosion on other environmental processes at the affected sites can often cause net CO <sub>2</sub> emissions through their indirect influence on soil fertility, and the balance of organic carbon inputs and outputs, interacting with other non-erosive soil degradation processes (such as nutrient depletion, compaction and salinisation), which can lead to the same net carbon effects (see Table 4.1) (van de Koppel et al. 1997 <sup>[[#fn:r786|786]]</sup> ).

As natural and human-induced erosion can result in net carbon storage in very stable buried pools at the deposition locations, degradation in those locations has a high C-release potential. Coastal ecosystems such as mangrove forests, marshes and seagrasses are at typical deposition locations, and their degradation or replacement with other vegetation is resulting in a substantial carbon release (0.15 to 1.02 GtC yr <sup>–1</sup> ) (Pendleton et al. 2012 <sup>[[#fn:r787|787]]</sup> ), which highlights the need for a spatially integrated assessment of land degradation impacts on climate that considers in-situ but also ex-situ emissions.

Cultivation and agricultural management of cultivated land are relevant in terms of global CO <sub>2</sub> land–atmosphere exchange (Section 4.8.1). Besides the initial pulse of CO <sub>2</sub> emissions associated with the onset of cultivation and associated vegetation clearing (Chapter 2), agricultural management practices can increase or reduce carbon losses to the atmosphere. Although global croplands are considered to be at a relatively neutral stage in the current decade (Houghton et al. 2012 <sup>[[#fn:r788|788]]</sup> ), this results from a highly uncertain balance between coexisting net losses and gains. Degradation losses of soil and biomass carbon appear to be compensated by gains from soil protection and restoration practices such as cover crops, conservation tillage and nutrient replenishment favouring organic matter build-up. Cover crops, increasingly used to improve soils, have the potential to sequester 0.12 GtC yr <sup>–1</sup> on global croplands with a saturation time of more than 150 years (Poeplau and Don 2015 <sup>[[#fn:r789|789]]</sup> ). No-till practices (i.e., tillage elimination favouring crop residue retention in the soil surface) which were implemented to protect soils from erosion and reduce land preparation times, were also seen with optimism as a carbon sequestration option, which today is considered more modest globally and, in some systems, even less certain (VandenBygaart 2016 <sup>[[#fn:r799|799]]</sup> ; Cheesman et al. 2016 <sup>[[#fn:r791|791]]</sup> ; Powlson et al. 2014 <sup>[[#fn:r792|792]]</sup> ). Among soil fertility restoration practices, lime application for acidity correction, increasingly important in tropical regions, can generate a significant net CO <sub>2</sub> source in some soils (Bernoux et al. 2003 <sup>[[#fn:r793|793]]</sup> ; Desalegn et al. 2017 <sup>[[#fn:r794|794]]</sup> ).

Land degradation processes in seminatural ecosystems driven by unsustainable uses of their vegetation through logging or grazing lead to reduced plant cover and biomass stocks, causing net carbon releases from soils and plant stocks. Degradation by logging activities is particularly prevalent in developing tropical and subtropical regions, involving carbon releases that exceed by far the biomass of harvested products, including additional vegetation and soil sources that are estimated to reach 0.6 GtC yr <sup>–1</sup> (Pearson et al. 2014, 2017 <sup>[[#fn:r795|795]]</sup> ). Excessive grazing pressures pose a more complex picture with variable magnitudes and even signs of carbon exchanges. A general trend of higher carbon losses in humid overgrazed rangelands suggests a high potential for carbon sequestration following the rehabilitation of those systems (Conant and Paustian 2002 <sup>[[#fn:r796|796]]</sup> ) with a global potential sequestration of 0.045 GtC yr <sup>-1</sup> . A special case of degradation in rangelands is the process leading to the woody encroachment of grass-dominated systems, which can be responsible for declining animal production but high carbon sequestration rates (Asner et al. 2003 <sup>[[#fn:r797|797]]</sup> ; Maestre et al. 2009 <sup>[[#fn:r798|798]]</sup> ).

Fire regime shifts in wild and seminatural ecosystems can become a degradation process in itself, with high impact on net carbon emission and with underlying interactive human and natural drivers such as burning policies (Van Wilgen et al. 2004 <sup>[[#fn:r1651|1651]]</sup> ), biological invasions (Brooks et al. 2009 <sup>[[#fn:r800|800]]</sup> ), and plant pest/disease spread (Kulakowski et al. 2003 <sup>[[#fn:r801|801]]</sup> ). Some of these interactive processes affecting unmanaged forests have resulted in massive carbon release, highlighting how degradation feedbacks on climate are not restricted to intensively used land but can affect wild ecosystems as well (Kurz et al. 2008 <sup>[[#fn:r802|802]]</sup> ).

Agricultural land and wetlands represent the dominant source of non-CO <sub>2</sub> greenhouse gases (GHGs) (Chen et al. 2018d <sup>[[#fn:r803|803]]</sup> ). In agricultural land, the expansion of rice cultivation (increasing CH <sub>4</sub> sources), ruminant stocks and manure disposal (increasing CH <sub>4</sub> , N <sub>2</sub> O and NH <sub>3</sub> fluxes) and nitrogen over-fertilisation combined with soil acidification (increasing N <sub>2</sub> O fluxes) are introducing the major impacts ( ''medium agreement, medium evidence'' ) and their associated emissions appear to be exacerbated by global warming ( ''medium agreement, medium evidence'' ) (Oertel et al. 2016 <sup>[[#fn:r804|804]]</sup> ).

As the major sources of global N <sub>2</sub> O emissions, over-fertilisation and manure disposal are not only increasing in-situ sources but also stimulating those along the pathway of dissolved inorganic nitrogen transport all the way from draining waters to the ocean ( ''high agreement, medium evidence'' ). Current budgets of anthropogenically fixed nitrogen on the Earth System (Tian et al. 2015 <sup>[[#fn:r805|805]]</sup> ; Schaefer et al. 2016 <sup>[[#fn:r806|806]]</sup> ; Wang et al. 2017a <sup>[[#fn:r807|807]]</sup> ) suggest that N <sub>2</sub> O release from terrestrial soils and wetlands accounts for 10–15% of the emissions, yet many further release fluxes along the hydrological pathway remain uncertain, with emissions from oceanic ‘dead-zones’ being a major aspect of concern (Schlesinger 2009; Rabalais et al. 2014 <sup>[[#fn:r808|808]]</sup> ).

Environmental degradation processes focused on the hydrological system, which are typically manifested at the landscape scale, include both drying (as in drained wetlands or lowlands) and wetting trends (as in waterlogged and flooded plains). Drying of wetlands reduces CH <sub>4</sub> emissions (Turetsky et al. 2014 <sup>[[#fn:r812|812]]</sup> ) but favours pulses of organic matter mineralisation linked to high N <sub>2</sub> O release (Morse and Bernhardt 2013 <sup>[[#fn:r813|813]]</sup> ; Norton et al. 2011 <sup>[[#fn:r814|814]]</sup> ). The net warming balance of these two effects is not resolved and may be strongly variable across different types of wetlands. In the case of flooding of non-wetland soils, a suppression of CO <sub>2</sub> release is typically overcompensated in terms of net greenhouse impact by enhanced CH <sub>4</sub> fluxes that stem from the lack of aeration but are aided by the direct effect of extreme wetting on the solubilisation and transport of organic substrates (McNicol and Silver 2014 <sup>[[#fn:r815|815]]</sup> ). Both wetlands rewetting/restoration and artificial wetland creation can increase CH <sub>4</sub> release (Altor and Mitsch 2006 <sup>[[#fn:r816|816]]</sup> ; Fenner et al. 2011 <sup>[[#fn:r817|817]]</sup> ). Permafrost thawing is another major source of CH <sub>4</sub> release, with substantial long-term contributions to the atmosphere that are starting to be globally quantified (Christensen et al. 2004 <sup>[[#fn:r818|818]]</sup> ; Schuur et al. 2015 <sup>[[#fn:r819|819]]</sup> ; Walter Anthony et al. 2016 <sup>[[#fn:r820|820]]</sup> ).

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