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

=== 4.9.4 Degradation and management of peat soils ===

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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> ).

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