Editing IPCC:AR6/WGII/Chapter-2 (section)

==== 2.4.3.4 Observed Changes in Mediterranean-Type Ecosystems ====

<div id="h3-18-siblings" class="h3-siblings"></div>

Since AR5 ( [[#Settele--2014|Settele et al. (2014)]] , all five Mediterranean-type ecosystems (MTEs) of the world have experienced extreme droughts within the past decade, with South Africa and California reporting their worst on record ( ''robust evidence'' , ''high agreement'' ) ( [[#Diffenbaugh--2015|Diffenbaugh et al., 2015]] ; [[#Williams--2015a|Williams et al., 2015a]] ; [[#Garreaud--2017|Garreaud et al., 2017]] ; [[#Otto--2018|Otto et al., 2018]] ; [[#Sousa--2018|Sousa et al., 2018]] ). Climate change is causing these droughts to become more frequent and severe ( ''medium evidence'' , ''medium agreement'' ) ( [[#AghaKouchak--2014|AghaKouchak et al., 2014]] ; [[#Garreaud--2017|Garreaud et al., 2017]] ; [[#Otto--2018|Otto et al., 2018]] ; [[#Seneviratne--2021|Seneviratne et al., 2021]] ).

MTEs show a range of direct responses to various forms of water deficit, but have also been affected by increasing fire activity linked to drought ( [[#Abatzoglou--2016|Abatzoglou and Williams, 2016]] ), and interactions between drought or extreme weather and fire affecting post-fire ecosystem recovery ( [[#Slingsby--2017|Slingsby et al., 2017]] ). Responses include shifts in functional composition ( [[#Acácio--2017|Acácio et al., 2017]] ; [[#Syphard--2019a|Syphard et al., 2019a]] ), decline of vegetation health ( [[#Hope--2014|Hope et al., 2014]] ; [[#Asner--2016a|Asner et al., 2016a]] ), decline or loss of characteristic species ( [[#White--2016|White et al., 2016]] ; [[#Stephenson--2019|Stephenson et al., 2019]] ), shifts in composition towards more drought- or heat-adapted species and declining diversity (see also section 2.4.4.3) ( [[#Slingsby--2017|Slingsby et al., 2017]] .; [[#Harrison--2018|Harrison et al., 2018]] ).

Declines in plant health and increased mortality in MTEs associated with drought have been widely documented ( ''robust evidence'' , ''high agreement'' ) ( [[#2.4.4.3|Section 2.4.4.3]] ). Remote-sensing studies show drought-associated mortality in post-fire vegetation regrowth in the Fynbos of South Africa ( [[#Slingsby--2020b|Slingsby et al., 2020b]] ), reduced canopy health in forests within MTE zones of South Africa ( [[#Hope--2014|Hope et al., 2014]] ) and declines in canopy water content in the forests of California ( [[#Asner--2016a|Asner et al., 2016a]] ). Several studies reported climate-associated responses of dominant or charismatic species. High mortality in the Clanwilliam cedar tree between 1931 and 2013 occurred at lower, hotter elevations in the Fynbos of South Africa ( [[#White--2016|White et al., 2016]] ). Drought reduced growth and increased mortality of the holm oak, ''Quercus ilex'' , on the Iberian Peninsula of Spain ( [[#Natalini--2016|Natalini et al. (2016)]] . Portuguese shrublands experienced losses of many deciduous and evergreen oak species, and an increasing dominance of pyrophytic xeric trees ( [[#Acácio--2017|Acácio et al., 2017]] ). The 2012–2015 drought in California caused high-canopy foliage dieback of the giant sequoia ( ''Sequoiadendron giganteum'' ) ( [[#Stephenson--2019|Stephenson et al., 2019]] ), increased the dominance of oaks relative to pines as a result of the increased water deficit, and led to large-scale tree mortality due to interactions of drought and insect pest outbreaks ( [[#McIntyre--2015|McIntyre et al., 2015]] ; [[#Fettig--2019|Fettig et al., 2019]] ).

Species distribution or community composition changes have contributed to declines in diversity and/or shifts towards more drought- or heat-adapted species ( ''medium evidence'' , ''high agreement'' ). Two conifer species ( ''Pinus longaeva'' and ''P. flexilis'' ) shifted upslope 19 m from 1950 to 2016 in the Great Basin, USA, ( [[#Smithers--2018|Smithers et al., 2018]] ). Reduced winter precipitation caused native annual forbs to recede, resulting in long-lasting and potentially unidirectional reductions in diversity in a Californian grassland ( [[#Harrison--2018|Harrison et al., 2018]] ). More frequent extreme hot and dry weather between 1966 and 2010 caused a decline in diversity during the post-fire regeneration phase in the Fynbos of South Africa ( [[#Slingsby--2017|Slingsby et al., 2017]] ), resulting in shifts towards species with higher temperature preferences ( [[#Slingsby--2017|Slingsby et al., 2017]] ). In Italy, [[#Del%20Vecchio--2015|Del Vecchio et al. (2015)]] observed increases in plant cover and thermophilic species in coastal foredune habitats between 1989 and 2012.

In southern California, USA, areas of forest and woody shrublands are shifting to grasslands, driven by a combination of climate and land use factors such as increased drought, fire ignition frequency and increases in nitrogen deposition ( ''robust evidence'' , ''high agreement'' ) ( [[#Jacobsen--2018|Jacobsen and Pratt, 2018]] ; [[#Park--2018|Park et al., 2018]] ; [[#Park--2019|Park and Jenerette, 2019]] ; [[#Syphard--2019b|Syphard et al., 2019b]] ).

The effects of climate change on heat, fuel and wildfire ignition limits show spatial and temporal variation globally (see [[#2.3|Section 2.3.6.1]] ), but there have been a number of observed impacts on MTEs ( ''medium evidence'' , ''high agreement'' ). Climate change caused increases in fuel aridity and the area of land burned by wildfires across the western USA from 1985 to 2015 ( [[#Abatzoglou--2016|Abatzoglou and Williams, 2016]] ). Local and global climatic variability led to a 4-year decrease in the average fire return time in the Fynbos, South Africa, when comparing fires recorded in 1951–1975 and 1976–2000 ( [[#Wilson--2010|Wilson et al., 2010]] ). In Chile, [[#González--2018|González et al. (2018)]] reported a significant increase in the number, size, duration and simultaneity of large fires during the 2010–2015 ‘megadrought’ when compared to the 1990–2009 baseline.

<div id="2.4.3.5" class="h3-container"></div>

<span id="observed-changes-in-savanna-and-grasslands"></span>