Editing IPCC:AR6/SRCCL/Chapter-3 (section)

==== 3.6.1.2 Grazing and fire management in drylands ====

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Rangeland management systems such as sustainable grazing approaches and re-vegetation increase rangeland productivity ( ''high confidence'' ) (Table 6.5). Open grassland, savannah and woodland are home to the majority of world’s livestock production (Safriel et al. 2005 <sup>[[#fn:r1029|1029]]</sup> ). Within these drylands areas, prevailing grazing and fire regimes play an important role in shaping the relative abundance of trees versus grasses (Scholes and Archer 1997 <sup>[[#fn:r1030|1030]]</sup> ; Staver et al. 2011 <sup>[[#fn:r1031|1031]]</sup> ; Stevens et al. 2017 <sup>[[#fn:r1032|1032]]</sup> ), as well as the health of the grass layer in terms of primary production, species richness and basal cover (the propotion of the plant that is in the soil) (Plaza-Bonilla et al. 2015 <sup>[[#fn:r1033|1033]]</sup> ; Short et al. 2003 <sup>[[#fn:r1034|1034]]</sup> ). This in turn influences levels of soil erosion, soil nutrients, secondary production and additional ecosystem services (Divinsky et al. 2017 <sup>[[#fn:r1035|1035]]</sup> ; Pellegrini et al. 2017 <sup>[[#fn:r1036|1036]]</sup> ). A further set of drivers, including soil type, annual rainfall and changes in atmospheric CO <sub>2</sub> may also define observed rangeland structure and composition (Devine et al. 2017 <sup>[[#fn:r1037|1037]]</sup> ; Donohue et al. 2013 <sup>[[#fn:r1038|1038]]</sup> ), but the two principal factors that pastoralists can manage are grazing and fire, by altering their frequency, type and intensity.

The impact of grazing and fire regimes on biodiversity, soil nutrients, primary production and further ecosystem services is not constant and varies between locations (Divinsky et al. 2017 <sup>[[#fn:r1039|1039]]</sup> ; Fleischner 1994 <sup>[[#fn:r1040|1040]]</sup> ; van Oijen et al. 2018 <sup>[[#fn:r1041|1041]]</sup> ). Trade-offs may therefore need to be considered to ensure that rangeland diversity and production are resilient to climate change (Plaza-Bonilla et al. 2015 <sup>[[#fn:r1042|1042]]</sup> ; van Oijen et al. 2018 <sup>[[#fn:r1043|1043]]</sup> ). In certain locations, even light to moderate grazing has led to a significant decrease in the occurrence of particular species, especially forbs (O’Connor et al. 2011 <sup>[[#fn:r1044|1044]]</sup> ; Scott-shaw and Morris 2015 <sup>[[#fn:r1045|1045]]</sup> ). In other locations, species richness is only significantly impacted by heavy grazing and is able to withstand light to moderate grazing (Divinsky et al. 2017 <sup>[[#fn:r1046|1046]]</sup> ). A context specific evaluation of how grazing and fire impact particular species may therefore be required to ensure the persistence of target species over time (Marty 2005 <sup>[[#fn:r1047|1047]]</sup> ). A similar trade-off may need to be considered between soil carbon sequestration and livestock production. As noted by Plaza-Bonilla et al. (2015) <sup>[[#fn:r1048|1048]]</sup> increasing grazing pressure has been found to increase SOC stocks in some locations, and decrease them in others. Where it has led to a decrease in soil carbon stocks, for example in Mongolia (Han et al. 2008 <sup>[[#fn:r1049|1049]]</sup> ) and Ethiopia (Bikila et al. 2016 <sup>[[#fn:r1050|1050]]</sup> ), trade-offs between carbon sequestration and the value of livestock to local livelihoods need be considered.

Although certain herbaceous species may be unable to tolerate grazing pressure, a complete lack of grazing or fire may not be desired in terms of ecosystems health. It can lead to a decrease in basal cover and the accumulation of moribund, unpalatable biomass that inhibits primary production (Manson et al. 2007 <sup>[[#fn:r1051|1051]]</sup> ; Scholes 2009 <sup>[[#fn:r1052|1052]]</sup> ). The utilisation of the grass sward through light to moderate grazing stimulates the growth of biomass and basal cover, and allows water services to be sustained over time (Papanastasis et al. 2017 <sup>[[#fn:r1053|1053]]</sup> ; Scholes 2009 <sup>[[#fn:r1054|1054]]</sup> ). Even moderate to heavy grazing in periods of higher rainfall may be sustainable, but constant heavy grazing during dry periods, and especially droughts, can lead to a reduction in basal cover, SOC, biological soil crusts, ecosystem services and an accelerated erosion ( ''high agreement, robust evidence'' ) (Archer et al. 2017 <sup>[[#fn:r1055|1055]]</sup> ; Conant and Paustian 2003 <sup>[[#fn:r1056|1056]]</sup> ; D’Odorico et al. 2013 <sup>[[#fn:r1057|1057]]</sup> ; Geist and Lambin 2004 <sup>[[#fn:r1058|1058]]</sup> ; Havstad et al. 2006 <sup>[[#fn:r1059|1059]]</sup> ; Huang et al. 2007 <sup>[[#fn:r1060|1060]]</sup> ; Manzano and Návar 2000 <sup>[[#fn:r1061|1061]]</sup> ; Pointing and Belnap 2012 <sup>[[#fn:r1062|1062]]</sup> ; Weber et al. 2016 <sup>[[#fn:r1063|1063]]</sup> ). For this reason, the inclusion of drought forecasts and contingency planning in grazing and fire management programmes is crucial to avoid desertification (Smith and Foran 1992 <sup>[[#fn:r1064|1064]]</sup> ; Torell et al. 2010 <sup>[[#fn:r1065|1065]]</sup> ). It is an important component of avoiding and reducing early degradation. Although grasslands systems may be relatively resilient and can often recover from a moderately degraded state (Khishigbayar et al. 2015 <sup>[[#fn:r1066|1066]]</sup> ; Porensky et al. 2016 <sup>[[#fn:r1067|1067]]</sup> ), if a tipping point has been exceeded, restoration to a historic state may not be economical or ecologically feasible (D’Odorico et al. 2013 <sup>[[#fn:r1068|1068]]</sup> ).

Together with livestock management (Table 6.5), the use of fire is an integral part of rangeland management, which can be applied to remove moribund and unpalatable forage, exotic weeds and woody species (Archer et al. 2017 <sup>[[#fn:r1069|1069]]</sup> ). Fire has less of an effect on SOC and soil nutrients in comparison to grazing (Abril et al. 2005 <sup>[[#fn:r1070|1070]]</sup> ), yet elevated fire frequency has been observed to lead to a decrease in soil carbon and nitrogen (Abril et al. 2005 <sup>[[#fn:r1071|1071]]</sup> ; Bikila et al. 2016 <sup>[[#fn:r1072|1072]]</sup> ; Bird et al. 2000 <sup>[[#fn:r1073|1073]]</sup> ; Pellegrini et al. 2017 <sup>[[#fn:r1074|1074]]</sup> ). Although the impact of climate change on fire frequency and intensity may not be clear due to its differing impact on fuel accumulation, suitable weather conditions and sources of ignition (Abatzoglou et al. 2018 <sup>[[#fn:r1075|1075]]</sup> ; Littell et al. 2018 <sup>[[#fn:r1076|1076]]</sup> ; Moritz et al. 2012 <sup>[[#fn:r1077|1077]]</sup> ), there is an increasing use of prescribed fire to address several global change phenomena, for example, the spread of invasive species and bush encroachment, as well as the threat of intense runaway fires (Fernandes et al. 2013 <sup>[[#fn:r1078|1078]]</sup> ; McCaw 2013 <sup>[[#fn:r1079|1079]]</sup> ; van Wilgen et al. 2010 <sup>[[#fn:r1080|1080]]</sup> ). Cross-Chapter Box 3 in Chapter 2 provides a further review of the interaction between fire and climate change.

There is often much emphasis on reducing and reversing the degradation of rangelands due to the wealth of benefits they provide, especially in the context of assisting dryland communities to adapt to climate change (Webb et al. 2017 <sup>[[#fn:r1081|1081]]</sup> ; Woollen et al. 2016 <sup>[[#fn:r1082|1082]]</sup> ). The emerging concept of ecosystem-based adaptation has highlighted the broad range of important ecosystem services that healthy rangelands can provide in a resilient manner to local residents and downstream economies (Kloos and Renaud 2016 <sup>[[#fn:r1083|1083]]</sup> ; Reid et al. 2018 <sup>[[#fn:r1084|1084]]</sup> ). In terms of climate change mitigation, the contribution of rangelands, woodland and sub-humid dry forest (e.g., Miombo woodland in south-central Africa) is often undervalued due to relatively low carbon stocks per hectare. Yet due to their sheer extent, the amount of carbon sequestered in these ecosystems is substantial and can make a valuable contribution to climate change mitigation (Lal 2004 <sup>[[#fn:r1085|1085]]</sup> ; Pelletier et al. 2018 <sup>[[#fn:r1086|1086]]</sup> ).

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