INTRODUCTION
With increased international trade, tourism, and global climate change,
biological invasions by insect pests have become serious threats to both
agricultural and forest ecosystems (Brockerhoff & Liebhold 2017; Seidlet al . 2018). A global approach is needed to develop effective
risk
mitigation
and management strategies against invasive insect pests which requires
understanding not only the species’ ecology and invasion process in the
newly invaded habitat, but also factors affecting its distribution and
spatiotemporal dynamics in its native habitat (Herms & McCullough 2014;
Jones 2019).
Native to northeastern Asia (China, Russian Far East, Japan, and South
Korea), the emerald ash borer (EAB), Agrilus planipennisFairmaire, was accidentally introduced to North America in the 1990s via
cargo crates, dunnage, or wood pallets originating from China (Siegertet al . 2014). Since it was first detected in southern Michigan,
U.S., and Ontario, Canada in 2002 (Poland & McCullough 2006), this
beetle has now spread to 35 U.S. states and five Canadian provinces (EAB
Info 2020). In the process of this invasion, EAB has killed hundreds of
millions of North American ash trees (Fraxinus spp.). More
recently, this beetle has also been discovered attacking the white
fringe tree, Chionanthus virginicus L., native to southern U.S.
and commonly planted as an ornamental and landscape tree throughout the
region (Peterson & Cipollini 2017; USDA NRCS 2019a). In a recent
laboratory study, EAB also completed development on olive trees,Olea europaea L., under artificially-forced infestations
(Cipollini et al . 2017) and thus has the potential to cause
serious economic loss to olive crops in addition to
degradation of North American
hardwood forest ecosystems (CFIA 2019; USDA 2019; EAB Info 2020).
All species of North American ash appear susceptible to EAB (Anulewiczet al . 2008); however, preference and susceptibility vary among
species. Green ash (F. pennsylvanica Marsh.) trees were preferred
and had higher levels of canopy dieback and EAB densities compared to
white ash (F. americana L.) trees at the same sites, while white
ash was preferred over blue ash (F. quadrangulate Michx.) trees
(Anulewicz et al . 2007; Tanis & McCullough 2012). White and
green ash cultivars had higher levels of EAB attack density and
mortality than a Manchurian ash (F. mandshurica Rupr.) cultivar
planted in a common garden trial in southeast Michigan (Rebek et
al . 2008), suggesting that Asian ash species developed resistance
mechanisms through their evolutionary history with EAB that is lacking
with North American species (Rebek et al . 2008; Villari et
al . 2016). Differences in susceptibility to EAB among ash species may
be related to differences in host volatiles, nutrition, and defense
compounds (Eyles et al . 2007; Whitehill et al . 2012;
Cipollini et al . 2019). It was found that EAB adults preferred to
feed on green, white and black ash (F. nigra Marsh.) leaves
compared to European (F. excelsior L.), blue or Manchurian ash
leaves, and ash species differed significantly in the relative amounts
of
antennally-active
volatiles (Pureswaran & Poland 2009).
In China, EAB was an occasional but largely unnoticed pest of ash trees
until the 1960s when North American ash species became widely introduced
as plantation trees in northern China (Wei et al . 2004). EAB has
not been recorded attacking any host plants (includingChionanthus spp.) other than ash trees in China and elsewhere in
Asia (Valenta et al . 2017). Scientific literature on the biology
and ecology of EAB was sparse before the 1960s, and even up to the 1980s
there were only limited descriptions of its damage and life-history from
anecdotal observations in books and regional publications (Wei et
al . 2004). However, extensive field and laboratory studies of EAB and
its natural enemies in northeast Asia (particularly in China) were
conducted in the 2000s after this beetle became a serious invasive pest
in North America. Consequently, knowledge about EAB’s distribution,
biology and associated natural enemies in its native (Wang et al .
2010; Wang et al . 2015) and newly invaded ranges (Duan et
al . 2018) has greatly accumulated. In North America, EAB has
established populations in warmer climate zones (as far south as 32°N)
(McConnell et al . 2019; EAB Info 2020) than in its native range
in China (as far south as 36°N) (Orlova-Bienkowskaja & Volkovitsh
2018). The cause of EAB outbreaks in China, and the apparent preference
for different climate zones in its invaded range are currently unknown.
In the present study, we conducted field surveys from 2003 to 2019 and
reviewed associated literature
published since 1900 to gather and analyze historical data on
occurrence, distribution, and outbreaks of EAB in China, and examine the
role of host plants (ash species) in influencing the spatiotemporal
dynamics of EAB’s occurrence or outbreaks. Through analyzing historical
data along with field observations of EAB occurrence and host and site
characteristics, we aimed to determine the causes of EAB outbreaks and
its geographic distribution in its native range. Findings from the
present study should improve our understanding of EAB’s geographic
distribution in both its native and invaded ranges, and may contribute
to development of sustainable management strategies against EAB.