1 INTRODUCTION
Fire plays a key role in ecosystems across the Earth, influencing
species composition, physical structure, and processes (Bond et al.,
2005; Krawchuk et al., 2009; Pausas & Keeley, 2009; Archibald et al.,
2018; Bowman et al. 2020). Natural disturbances such as fire have long
been recognized as regulators of biological diversity (Connell, 1978;
Huston, 1994). Understanding spatial and temporal variation in the
fire-diversity nexus is especially critical given the profound
anthropogenic alterations of fire regimes across the Earth and their
cascading impacts on ecosystems and species, including humans (Bowman et
al., 2020; Coop et al., 2020). Ecologists have responded by intensifying
their efforts to develop generalizations about species diversity and
fire that address the challenges of a fierier world in the Anthropocene
(e.g., Perry et al., 2011; Burkle et al., 2015; Enright et al., 2015;
Pausas & Ribiero, 2017; He et al., 2019; Bowman et al., 2020; Coop et
al., 2020; Miller & Safford, 2020)
Whittaker (1970, 1972) partitioned species diversity into three
components: alpha diversity (α), beta diversity (β), and gamma diversity
(γ). α is species diversity at a point in the landscape (i.e., a site),
which itself can be decomposed into the number of species (richness) and
the evenness of abundances among species. α is usually measured as the
mean or median diversity of multiple local sites distributed across an
area of study. β captures differences in species assemblages from one
site to another and has been measured with a wide variety of approaches
(Anderson et al., 2011). The combination of α and β produces γ, the
total species diversity supported in the larger area—landscape
diversity. Although myriad hypotheses have been proposed and tested
regarding the relationship between fire and these three levels of
diversity, generalizations have been elusive, which signals the need for
further conceptual work and hypothesis testing for understanding the
variety of ways in which fire influences diversity (Parr & Andersen,
2006; Anderson et al., 2014; Burkle et al., 2015; Kelly & Brotons,
2017; He et al., 2019; Miller & Safford, 2020). In this paper, we test
key hypotheses on the impact of wildfire on woody plant diversity in a
topographically complex mountain range.
The intermediate disturbance hypothesis (IDH) proposes that species
richness—usually α, less commonly β and γ—varies predictably with
disturbance gradients in a unimodal, hump-shaped fashion, in which
intermediate levels of disturbance intensity or frequency maintain high
diversity (Connell, 1978; Sousa, 1979). The IDH has been vigorously
debated (Fox, 2013; Huston, 2014), with mixed support across a wide
range of disturbance types and taxa (Sheil & Burslem, 2013).
Nevertheless, research has repeatedly revealed a hump-shaped
relationship between plant species richness and fire severity (DeSiervo
et al., 2015; He et al., 2019; Richter et al. 2019; Strand et al. 2019;
Miller & Safford, 2020), especially in frequent, low severity fire
regimes (Miller & Safford, 2020). The assumed underlying mechanisms
vary, but most propose that different fire severities environmentally
select for different sets of species. Under no fire or low fire
severity, for example, competitive and fire-resistant species should
thrive, whereas, after a high severity fire, fast growing, rapidly
colonizing species should predominate. The IDH proposes that at
intermediate fire severity, both sets of species can coexist, resulting
in a peak in species richness. Despite support for the IDH for fire,
other studies have detected neutral, linear positive, and negative
relationships between species richness and fire severity (He et al.,
2019; Miller & Safford, 2020).
Martin & Sapsis (1992) coined the term “pyrodiversity” to capture a
growing awareness of the ecological importance of variation across
landscapes in fire severity, frequency, size, and other attributes (see
also Krawchuk & Moritz, 2011; Perry et al., 2011; Bowman et al., 2016,
He et al., 2019). They argued that pyrodiversity promotes variation in
plant assemblages among sites (i.e,, β) because, as with the underlying
assumption of the IDH, different sets of species thrive under different
conditions related to fire, a phenomenon observed for decades in fire
prone ecosystems (Romme, 1982; Bond et al., 2005; Pausas & Ribiero,
2017). Mixed-severity fire regimes, for example, provide a complex
mosaic of post-fire conditions that should support a wider range of
plant species across a landscape than would low severity or high
severity fires alone. These arguments promoted an emerging management
dictum that prescribed burning aimed at fostering biodiversity (β and γ)
should create a broad spectrum of fire patch characteristics to provide
conditions required for the regeneration and persistence of a diverse
range of native biota (Perry et al., 2011; Bowman et al., 2016; Kelly &
Brotons, 2017). Adding patches of fire to an otherwise long unburned but
fire-prone area will generally enhance the diversity of most taxa (He et
al, 2019), but there’s disagreement about the rigor of field studies,
the shape of the relationship between fire and biodiversity, support for
underlying assumptions, and the strength of the evidence for broadly
applying these ideas to land management (Parr and Anderson, 2006; Perry
et al., 2011; Bowman et al., 2016; Kelly & Brotons, 2017).
Miller & Safford (2020) argue that the IDH and the pyrodiversity
hypotheses largely ignore the interaction of life history traits and the
historic fire regime of particular ecosystems. They propose that the
historic fire regime acts as a filter, selecting only those species with
the capacity to regenerate and persist under those conditions. Such life
history traits shaped over evolutionary time to adapt to the prevailing
fire regime are unlikely to confer similar success to fire regimes other
than the historic one. As an example, fire resistant tree species, with
thick insulative bark and regeneration from seed, perform well and often
dominate under frequent, surface fire regimes, but are readily killed in
ecosystems with infrequent, stand replacing fires (Barton & Poulos,
2018; Coop et al., 2020). This leads to the hypotheses that α richness
and β should peak at the historic fire severity of an ecosystem rather
than necessarily at intermediate severity as predicted by the IDH. It
predicts further that adding patches of fire outside of the historic
fire regime will not necessarily promote variation in species
assemblages across sites (i.e., β) or total landscape diversity (γ)
(Miller & Safford, 2020).
In an effort to further explore the relationships between landscape
variation in fire severity and woody plant species diversity, we
examined the impact of a 2011 megafire on α, β, and γ in an arid,
fire-prone mosaic of shrub, woodland, and forest ecosystems in
Chiricahua National Monument in the Sky Islands of Arizona, USA. Before
Euro-American settlement (<1880), frequent surface fires
predominated in conifer and conifer-oak forests (Swetnam et al., 1989;
Kaib et al., 1996; Swetnam & Baisan, 1996; Barton et al., 2001 Swetnam
et al., 2001), whereas more arid woodlands and interior chaparral
experienced mixed fire regimes with longer fire intervals (Kaib et al.,
1996; Baisan & Morino, 2000; Taylor et al., in press). Starting in the
late 1800s, reduction of fine fuel by livestock grazing and then active
suppression largely excluded fire for more than a century (Leopold,
1924; Marshall, 1957; Swetnam et al., 2001), until the 2011 Horseshoe
Two Megafire, which burned ~90,000 ha across the entire
mountain range, a size unprecedented in the historical fire record. This
fire was sparked by more than a century of mounting fuel loads and an
increasingly warmer and drier climate, part of a regional surge in very
large fires with a significant high-severity component throughout the
Southwest USA (Dennison et al., 2014; Abatzoglou & Williams, 2016;
Westerling, 2016; Singleton et al., 2019).
The Horseshoe Two Megafire offered the opportunity to evaluate temporal
shifts in woody plant diversity across a range of fire severities,
spanning from unburned to high-severity wildfire. To this end, we
sampled woody plant diversity in 138 plots before (2002-2003) and after
(2017-2018) the 2011 fire in three vegetation types and spanning wide
fire severity and topographic gradients. We specifically addressed (1)
whether α, β, and γ changed from the pre- to post-fire sample periods,
(2) the extent to which these changes were driven by the Horseshoe Two
Fire, (3) the direction and shape of the relationship of α, β, and γ to
fire severity and fire variability among plots, (4) whether diversity
patterns with respect to fire were tied to the underlying historic fire
regimes of the three vegetation types, and (5) the role of topography in
shaping biodiversity independent of the Horseshoe Two Fire.