Introduction

The emerald ash borer (EAB), Agrilus planipennis Fairmaire (Coleoptera: Buprestidae), native to Asia, is a destructive pest of ash trees, Fraxinus spp. (Oleaceae). This beetle was first discovered in southern Michigan, USA and nearby Ontario, Canada in 2002 and has since become the most destructive pest of North American ash-dominated hardwood forests (Herms and McCullough 2014). Attempts to eradicate EAB in North America by USA and Canadian regulatory agencies were abandoned a few years after its first detection because EAB became too widespread (GAO 2006). Ongoing research, development, and implementation of EAB-management strategies were subsequently directed towards management via biological control, regulatory restriction of movement of EAB-infested wood or plant materials, insecticide treatment or physical destruction of infested trees, and EAB-resistant ash genotypes (e.g., Liu et al. 2003, 2007; Bauer et al. 2015; Koch et al. 2015; McCullough et al. 2015; Mercader et al. 2015).

In 2020, the USA federal regulatory effort to contain the spread of EAB was discontinued because of high implementation costs, a lack of effective EAB-surveillance tools, and the inability to prevent EAB from both short-distance natural dispersal or long-distance spread by human transport of infested ash materials (Federal Register 2020). Although highly effective systemic insecticides are available to protect high value, landscape ash (Herms et al. 2009; Sadof et al. 2021), costs and environmental concerns prevent widespread use of chemical controls against EAB in natural forests. Therefore, classical biological control via the discovery, introduction, release, and establishment of self-propagating and dispersing host-specific natural enemies is currently the most promising strategy for sustainable management of EAB to conserve native Fraxinus spp. in the forests of North America. Previous reviews by Bauer et al. (2015) and Duan et al. (2018) reported the progress in earlier phases of the classical biocontrol program against EAB in North America, including foreign exploration for Asian parasitoids, host specificity testing and regulatory approval of discovered agents, as well as release and establishment recovery in the USA and Canada. Here, we examine the ecological premises of protecting North American ash trees against EAB by the introduced natural enemies, with a particular focus on the population dynamics of the pest and introduced agents and implications for protecting and conserving native Fraxinus species in the aftermath of EAB invasion. In addition, we review recent progress and challenges in the implementation and evaluation of the biological control program against EAB as the target pest continues to spread throughout North America.

For this review, we first searched online data bases of both Scopus and ISI Web of Science using the key words “emerald ash borer or Agrilus planipennis” and found that both databases produced similar number of documents (904–922). Then, we conducted further searches of the Scopus online database using key words “emerald ash borer” or “Agrilus planipennis” in combination with “parasitoid”, “natural enemies”, “biological control” or “biocontrol”. A total of 228 unique publications were found that focused on biological control or biocontrol of the emerald ash borer with parasitoids or other natural enemies from the Scopus records. All these publications were read and analyzed for data and content related to the objectives of this review, namely ecological data on ash protection and conservation, and/or data on population dynamics of EAB and natural enemies. Data from relevant original research articles and significant reviews that addressed these objectives were used and cited in this review. In addition, we also contacted colleagues in China for any historical Chinese literatures that might be relevant to this review.

Ecological premise for protecting North American ash trees using classical biocontrol

Ash tree mortality risk from the EAB invasion

Ash, Fraxinus spp. (Oleaceae) trees were relatively free of serious diseases and insect pests in North America before the invasion of EAB (Pugh et al. 2011). Since it was first identified as the sole factor causing ash tree mortality in southeastern Michigan and nearby Ontario in 2002, however, the spread and establishment of EAB has killed hundreds of millions of North American ash trees in 36 states and Washington, D.C. in USA and five Canadian provinces (Canadian Food Inspection Agency 2022; Emerald Ash Borer Information 2022). The potential economic costs associated with the EAB invasion were estimated to be $1 billion per year from 2009 to 2019 (Kovacs et al. 2010), and the ecological impacts on North American forests are severe and widespread, including threats to vertebrate and invertebrate ash specialist herbivores (e.g., Wagner and Todd 2016); composition of arthropod communities associated with ash (Jennings et al. 2017); altered forest composition and structure (Morin et al. 2017); impacts on soil microbial communities, understory vegetation, and invasive plants (Klooster et al. 2018); altered riparian forest structure and function (Engelken and McCullough 2020; Engelken et. al. 2020); impacts on tidal swamps and coastal river habitat (Jacobsen 2020); and facilitation of secondary invasions (Baron and Rubin 2021).

All ash species native to North America encountered by EAB to date appear to be susceptible (Anulewicz et al. 2008; Herms and McCullough 2014), including the most common species: green (Fraxinus pensylvanica Marsh.), white (F. americana L.), and black (F. nigra Marsh.) as well as the less common blue (F. quadrangulata Michx.) and pumpkin ash (F. profunda [Bush] Bush). However, the degree of ash susceptibility to EAB varies among ash species and may be related to differences in bark texture, host volatiles, and nutritional or defensive compounds (e.g., Chen et al. 2011a, 2011b; Cipollini et al. 2011; Whitehill et al. 2011, 2012; Tanis and McCullough 2012, 2015; Koch et al. 2015; Rigsby et al. 2015; Villari et al. 2016; Qazi et al. 2018). Other ecological factors such as tree age, physiological condition, habitat type, and natural enemies may also play a role in ash susceptibility to EAB (Tluczek et al. 2011; Knight et al. 2014; Duan et al. 2021). For example, blue ash appears to be much less susceptible to EAB infestation than other North American ash trees, possibly due to a combination of differences in volatile emissions (Pureswaran and Poland 2009) and defense compounds (Qazi et al. 2018), as well as its smooth bark, which makes the tree less suitable for EAB oviposition (Tanis and McCullough 2012, 2015; Spei and Kashian 2017).

Although at some sites nearly 100% of ash trees > 2.5 cm in diameter at breast height (DBH) in infested stands have been attacked and killed by EAB (Klooster et al. 2014), invading EAB populations in North America appear to first kill mature (canopy) ash trees as compared to smaller understory ash trees, saplings, and seedlings (Cappeart et al. 2005; Tanis and McCullough 2015). Moreover, smaller ash saplings with DBH < 2.5 cm are rarely attacked by EAB (Marshall et al. 2013). It is conceivable that younger ash trees have both physical (e.g., smooth-bark surface) and chemical (secondary defense compounds) characteristics that are less attractive to EAB oviposition than canopy ash trees (e.g., Marshall et al. 2013). It is also possible that the smaller area of phloem in these young ash trees limits EAB colonization.

Klooster et al. (2014) found that the three most common North American ash species, green, white, and black, are equally vulnerable to severe levels of mortality when EAB populations are high. To date, however, few studies have determined the threshold EAB density that kills an ash tree of a given species at a given age in a specific habitat. A study in urban forests in Canada suggests that infested ash trees could recover from a density of 10 EAB larvae per m2 of phloem (MacQuarrie and Scharbach 2015). A more recent study in Michigan reported on the abundance of surviving ash saplings and young trees (DBH ~ 2.5–5.8 cm) in natural forests, where EAB densities averaged 2–7 larvae per m2 of phloem (Duan et al. 2017). It is plausible that the threshold EAB density that kills an ash tree varies with levels of tree resistance or tolerance, which are themselves influenced by a whole host of factors such as tree species, age, climates, and forest habitat conditions (e.g., MacQuarrie and Sabarbach 2015; Tennis and McCullough 2015; Dang et al. 2021).

Rationale for classical biocontrol

In contrast to the enormous economic and ecological impacts of EAB in North America, early Chinese literature only reported occasional damage to stressed or weakened Asian ash trees or susceptible North American ash trees in China (CASIZ 1986). Although at the time little was known about the biology of EAB and factors regulating its populations in Asia, the ability of North American ash trees to survive without significant mortality in China strongly suggested the possibility of effective top down EAB population control by specialized natural enemies. Subsequent discovery of a complex of hymenopteran parasitoids attacking EAB eggs and larvae in northern China and the Russian Far East further supports this premise of top-down suppression of EAB populations by the co-evolved natural enemies in its native range (see next Section).

Shortly after detection of EAB in North America, field surveys for native North American natural enemies attacking EAB were conducted in Michigan and other newly infested regions. Parasitism by native North American parasitoids was minimal (< 5%) in these invaded regions (e.g., see reviews in Bauer et al. 2015 and Davidson and Rieske 2016; Jennings et al. 2016; Duan et al. 2018). Relatively high levels of larval parasitism by generalist North American native parasitoids were observed at some heavily infested sites in Michigan and Ontario, with up to 71% parasitism by Atanycolus cappaerti Marsh and Strazanac (Hymenoptera: Braconidae) (Cappaert and McCullough 2009) and ~ 40% parasitism by Phasgonophora sulcata Westwood (Hymenoptera: Chalcididae) (Roscoe et al. 2016). Augmentative releases of the native natural enemy P. sulcata are also currently under investigation in Canada (Gaudon and Smith 2020). To date, however, no studies have demonstrated the effectiveness of these native North American parasitoids in regulating EAB population dynamics at low densities. Furthermore, no native North American parasitoids have been found attacking EAB eggs (Bauer et al. 2015; Duan et al. 2018), justifying exploration for co-evolved Asian natural enemies for implementation of classical biocontrol.

Implementation of an emerald ash borer biocontrol program

Bauer et al. (2015) and Duan et al. (2018) reviewed the progress in earlier phases of the classical biocontrol program against EAB in North America. Briefly, foreign explorations in the pest’s native range were conducted from 2003 to 2012 and led to discovery of three major hymenopteran parasitoids, an egg parasitoid Oobius agrili Zhang and Huang (Encyrtidae) (Zhang et al. 2005) and two larval parasitoids, Tetrastichus planipennisi Yang (Eulophidae) and Spathius agrili Yang (Braconidae) (Yang et al. 2005, 2006; Liu et al. 2007) in northeast China. While surveys for EAB natural enemies in Japan and Mongolia were unproductive because of the lack of detectable EAB populations, field work in the Russian Far East later resulted in discovery of an egg parasitoid Oobius primorskyensis Yao and Duan (Encyrtidae) and two larval parasitoids (Braconidae), Spathius galinae Belokobylskij & Strazanac and Atanycolus nigriventris Vojnovskaja-Krieger (Belokobylskij et al. 2012; Duan et al. 2012a; Yao et al. 2016). Additional exploration work in South Korea in areas with low density EAB populations infesting species of Asiatic ash in Daejeon and in Yangsuri recovered three natural enemy species including S. galinae, Tetrastichus telon Graham (Eulophidae), and a beetle Teneroides maculicollis Lewis (Cleridae) (Gould et al. 2015).

After reviewing host range data generated from both quarantine testing in the USA and China (Table 1), USDA APHIS issued environmental release permits for O. agrili, S. agrili, and T. planipennisi in 2007 and later for S. galinae in 2015. To produce large number of these parasitoids for environmental releases throughout EAB-infested regions in the USA, a mass-rearing facility was subsequently constructed by 2010 in Brighton, Michigan and field release guidelines for the introduced biocontrol agents were published (Gould et al. 2015; Duan et al. 2018; USDA–APHIS/ARS/FS 2021). The Brighton EAB biocontrol rearing facility currently produces ~ 400,000 female T. planipennisi, ~ 170,000 O. agrili, and ~ 100,000 female S. galinae annually. However, production of S. agrili was purposely reduced because releases were discontinued in northern regions after 2012 due to its lack of establishment at higher latitudes (USDA–APHIS/ARS/FS 2021). To improve the likelihood of establishment for both T. planipennisi and S. galinae, high numbers are being released at more northerly sites and at higher elevations where their synchrony with EAB larval host stages is confirmed (Gould et al. 2020; USDA–APHIS/ARS/FS 2021). By fall of 2022, one or more of these four biocontrol agents were released in > 360 counties in 31 EAB-infested states, Washington D.C., and three Canadian provinces (Mapbiocontrol 2022; Supplementary figure S1, S2, and S3; Butler et al. 2022).

Table 1 Non-target insect taxa tested with the Asian parasitoids petitioned for environmental release in North America as biocontrol agents against emerald ash borer, Agrilus planipennis (adapted from Duan et al. 2018)
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After parasitoid releases were made, establishment and spread of the released agents were evaluated with various sampling methods in the field, including ash tree-harvesting and rearing of parasitoids from large bolts (Butler et al. 2022), debarking of EAB-infested ash trees (Duan et al. 2013; Jennings et al. 2013), field deployment of sentinel EAB larvae and/or sentinel eggs as traps (Jennings et al. 2018; Rutledge et al. 2021; Quinn et al. 2022a, b), and use of yellow pan traps (Parisio et al. 2017; Petrice et al. 2021). Field recovery efforts revealed that all three of the biocontrol agents from China (O. agrili, T. planipennisi, and S. agrili) were recovered from EAB one year after release, indicating reproductive and overwintering success. However, only O. agrili and T. planipennisi were consistently recovered for two or more years after their last release in an area, and these two species are now considered firmly established and spreading naturally beyond their initial release sites. Recent studies on S. galinae in Michigan, Connecticut, Massachusetts, and New York, USA, where it was released from 2015 to 2017, have also shown that S. galinae is well established and spreading widely (Duan et al. 2019b, 2020; Rutledge et al. 2021; Quinn et al. 2022a). The establishment of S. agrili (i.e., recovery two years after the final release) is currently unconfirmed in northern states and most of the mid-Atlantic region. However, its reproduction in EAB larvae was confirmed one or two years after releases at a few sites south of the 40th parallel, where this species is still being released (Hooie et al. 2015; Jennings et al. 2016; Aker et al. 2022). Spathius agrili has not yet established well at any locations.

Several parasitoid recovery studies in the northern USA have documented rapid long-distance spread of T. planipennis and S. galinae following releases. For example, Jones et al. (2019) captured T. planipennisi in yellow pan traps deployed along the entirety of a 20 km transect in New York for three years following parasitoid release in a localized EAB outbreak. As EAB spread south along the ash corridor, T. planipennisi populations followed. Tetrastichus planipennisi in Michigan was found to have spread up to 3 km from release sites one year after its field releases (Duan et al. 2013). Using sentinel green ash logs infested with EAB larvae, Quinn et al. (2022a) detected both T. planipennisi and S. galinae 14 km away from the release sites 3–4 years after their last field releases in New York and Connecticut. Most recently, Aker et al. (2022) detected multiple established populations of S. galinae in Maryland at sites up to 90 km from the nearest release point approximately three year after release, indicating rapid, long-distance spread. In contrast to T. planipennisi and S. galinae, the egg parasitoid O. agrili appears to spread much more slowly in forests possibly because of its smaller size (~ 1 mm) and lower reproduction potential. For example, Abell et al. (2014) did not detect O. agrili at the non-release (control) sites 1.5–3 km away from the closest release sites until three years after the parasitoid’s release. Most recently, Quinn et al. (2022b) studied the dispersal ability of O. agrili by attaching freshly laid EAB eggs (sentinel hosts) on both green ash and white fringe trees located at various distances from the release point. Adult O. agrili were recovered at least 45 m from the release point within 4–5 days, and the dispersal distance was affected only by the time after initial release and not by hosts’ food plant species.

Evaluation of impacts of biocontrol on emerald ash borer and ash regeneration

Impact of introduced parasitoids on EAB population dynamics

Following environmental releases of T. planipennisi, O. agrili, and S. agrili, six long-term study sites, each comprised of a release and non-release control plot, were established in either 2007 or 2010 in southern Michigan, USA to monitor EAB population dynamics and mortality factors. At each release plot, ~ 1000–3000 female adults each of O. agrili, S. agrili, and T. planipennisi were released, and, in subsequent years, infested ash trees were sampled to estimate EAB egg and larval parasitism, and other causes of larval mortality (for details see Duan et al. 2013; Abell et al. 2014). Starting in 2015 after APHIS issued environmental release permits, S. galinae was also released for two consecutive years at each of these Michigan sites (Mapbiocontrol 2022).

EAB egg parasitism by O. agrili increased over the first five years after release at these Michigan sites, averaging ~ 1 to 4% from 2008 to 2011 and ~ 28% by 2014 in release plots. The natural spread of O. agrili from the release plots to the control plots occurred but was generally slow (Abell et al. 2014). The overall spread of O. agrili and its impact in suppressing EAB population growth have yet to be determined because sampling EAB eggs from ash bark layers and crevices is labor intensive and difficult to standardize (Abell et al. 2014; Petrice et al. 2021). Moreover, parasitism of EAB eggs by O. agrili is patchy. Therefore, more intensive sampling is needed to recover it and quantify its impact on EAB population dynamics (Petrice et al. 2021).

EAB larval parasitism by T. planipennisi in the Michigan plots was also low at first, averaging 1 to 6% from 2008 to 2011, but then increased to ~ 30% by 2014 in both the release and control plots (Duan et al. 2012b, 2013, 2015b). Life table analyses of seven years of data from the six Michigan study sites revealed that T. planipennisi contributed significantly to reducing net EAB population growth rates in smaller diameter trees in the aftermath of the initial EAB outbreak (Duan et al. 2015b). During the initial outbreak phase, native generalist natural enemies including parasitoids (Atanycolus spp.) and woodpeckers (such as Dryobates pubescens L., Leuconotopicus villosus L. and/or Melanerpes carolinus L.) contributed to declines in invasive EAB populations (Duan et al. 2014; Jennings et al. 2016). However, it was the introduced specialist T. planipennisi that became the dominant source of EAB larval mortality in small ash trees in the aftermath of the EAB invasion in Michigan (Duan et al. 2015b; 2017). A similar study in white ash forests of New York showed that the combination of woodpecker predation and parasitism by T. planipennisi significantly reduced the net reproductive rate of EAB in regenerating ash trees. At six sites in western New York the net reproductive rate was reduced to zero (Gould et al. 2022).

EAB larval parasitism by T. planipennisi was found to be concentrated in smaller diameter ash trees in field surveys in China, the Russian Far East, and the USA (Liu et al. 2007; Abell et al. 2012; Duan et al. 2012a; Jennings et al. 2016). The ability of T. planipennisi to parasitize EAB in larger ash trees is constrained by its short ovipositor (average 2 to 2.5 mm long), which cannot reach EAB larvae feeding under the thick bark (> 3.2 mm) found on the lower boles of ash trees that are > 12 cm DBH (Abell et al. 2012). Thus, T. planipennisi has a greater impact on EAB larval mortality in small diameter trees.

While releases of T. planipennisi and O. agrili are continuing, efforts are now increasingly focused on establishing S. agrili and S. galinae in North America to control EAB more successfully in larger ash trees, where their longer ovipositors allow them to reach EAB larvae under thicker bark. For example, the ovipositor of S. galinae averages 4 to 6 mm long, more than twice that of T. planipennisi. Consequently, S. galinae can attack EAB larvae feeding in ash trees up to 30 cm DBH (Duan et al. 2012a; Murphy et al. 2017). Based on their native distributions, S. agrili is presumed to be more adapted to southern climates and S. galinae to more northern climates (Jones et al. 2020). However, the lack of persistent recoveries of S. agrili from many of the previous release sites in Michigan (Duan et al. 2013, 2015b), New York and Maryland (Jennings et al. 2016), Tennessee (Hooie et al. 2015), and Kentucky (Davidson and Rieske 2016) suggests that this parasitoid may not be well adapted to North American hardwood forests for EAB biocontrol. In contrast, releases of S. galinae that began in the summer of 2015, primarily in Michigan and several Northeastern and Mid-Atlantic states, have resulted in successful establishment. Recent field studies from both Michigan and several northeastern USA states showed that S. galinae has established self-sustaining populations in release areas where T. planipennis had already been released or established (Duan et al. 2021, 2022). Based on recent life-table analyses, S. galinae alone caused a 31–57% reduction in the net population growth rates of EAB during the outbreak phase (Duan et al. 2022). Spathius galinae has now become the dominant parasitoid species, and, along with local generalist natural enemies and T. planipennisi, it reduced average EAB larval densities from 30 live EAB larvae per m2 of tree phloem in 2015 to less than seven in 2020 (Duan et al. 2022). This level of reduction has contributed to ash recovery and regeneration in the aftermath of EAB invasion waves (Duan et al. 2022). Life table analysis of EAB population dynamics at these biocontrol study sites indicates that the net population growth rate of EAB was at or below replacement levels (Figs. 1 and 2) (Duan et al. 2017, 2022).

Fig. 1
figure 1

Net population growth rates (R0) of emerald ash borer (Agrilus planipennis) infesting ash saplings (diameter at breast height or DBH = 2.5–5.8 cm), averaged from six different study sites in southern Michigan where the introduced larval parasitoids are well established since their releases from 2007 to 2010. Solid line represents R0 estimated using life table analysis by including all sources of the observed larval mortalities. Dashed line represents R0 estimated by the same lifetable analysis after excluding T. planipennisi from the life table, assuming mortality rates from other factors would not change due to increases in EAB densities (Duan et al. 2017)

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Fig. 2
figure 2

Net population growth rates (R0) of emerald ash borer (Agrilus planipennis) infesting pole-size ash trees (diameter at breast height or DBH = 8–27 cm), averaged across different study sites in northeastern states (Connecticut, New York, and Massachusetts), where the introduced larval parasitoid Spathius galinae is well established since its release in 2016 and 2017. Arrows indicate the timing of S. galinae releases: the small arrow represents low release numbers and the large arrow high release numbers. Solid line represents R0 estimated using life table analysis of the observed EAB larval survival and mortalities caused by S. galinae and other mortality factors (e.g., woodpeckers, other native parasitoids). Dashed line represents R0 estimated by the same lifetable analysis after removing parasitism by S. galinae from the life table, assuming mortality rates from other factors would not change due to increases in EAB densities (Duan et al. 2022)

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Recent studies suggest that the success or effectiveness of the current EAB biocontrol program in North America may be influenced by the interaction of EAB and parasitoid lifecycles. Jones et al. (2020) found that S. galinae and T. planipennisi are well synchronized with a lifecycle where some EAB take two years to develop as is the case in the northern USA. However, in more southern states, where most EAB overwinter as mature larvae in pupation chambers out of reach of parasitoids emerging in early spring, parasitoid populations are less likely to persist. In fact, Gould et al. (2020) modeled the likelihood of establishment of T. planipennisi and found that as summer temperatures increased, the percentage of EAB overwintering under the bark and thus available to spring emerging parasitoids decreased, and the likelihood of establishment by T. planipennisi also declined. Spathius galinae, unlike S. agrili which emerges in mid-summer (Yang et al. 2010), also emerges early in the spring (Jones et al. 2020) and is likely to establish more poorly in southern states. Oobius agrili has two generations per year. The first generation emerges from diapause and produces the second generation progenies that are able to reproduce without requiring diapause. Non-diapause adults of the second generation reproduces and their offspring enters winter diapause. Petrice et al. (2019) found that O. agrili has a critical daylength threshold for entering diapause, and that predicted synchrony with EAB egg laying is affected by latitude and thus daylength throughout the summer. Oobius primorskyensis, which has different diapause requirements than O. agrili, might be able to survive and thrive where O. agrili does not establish (Larson and Duan 2016; Duan et al. 2019a).

Ash recovery and regeneration in the aftermath of EAB invasion with biocontrol

Because the high abundance of susceptible North American ash species facilitates rapid EAB population growth rates, it would be extremely difficult to rapidly protect susceptible overstory ash trees against EAB in newly invaded-areas solely through the introduction and establishment of limited numbers of specialized natural enemies from Asia. In post-EAB invaded hardwood forests of North America, however, where EAB populations are lower and ash densities have been dramatically reduced, establishment of the introduced EAB parasitoids may effectively conserve surviving ash by moderating the frequency and amplitude of future EAB outbreaks, as occurs in EAB’s native range (see previous section). This in turn should allow these surviving trees to increase in age and their reproduction should lead to higher ash densities over time.

In southeastern Michigan, where establishment and spread of T. planipennisi and O. agrili have been confirmed since 2012, densities of ash and other native saplings were higher in forests closer to parasitoid release sites (Margulies et al. 2017). In another study of ash health in long-term EAB biocontrol study sites in 2012 and 2015, lower tree mortality and greater diameter growth were observed in large diameter ash trees growing in release plots vs. those in control plots (Kashian et al. 2018). Moreover, researchers found that many relatively healthy ash saplings (4–16 per 100 m2) and pole-size young trees (2–9 per 100 m2) have persisted, despite formerly high EAB densities that resulted in loss of most overstory ash trees by 2010 (Duan et al. 2017; Gould et al. 2022; JJD, TRP unpublished data). However, recovery of North American ash in the post-EAB invasion forests will take time even after EAB densities are successfully reduced by the introduced agents. This is because tree regrowth and regeneration are very slow processes, normally taking more than two decades for these ash trees to reach the overstory.

Concluding remarks

Since its first detection in the USA in 2002, EAB has continued to spread and cause economic damage to ash nursery stock and the lumber industry, degradation of ash forests, and reduction in ecosystem functions in ash forests in North America. The classical biocontrol program against EAB, which started over a decade ago with the introduction and establishment of co-evolved natural enemies from the pest’s native range, has shown the ability to suppress EAB to lower densities, which is allowing North American ash species in northern hardwood forests to recover and regenerate in the aftermath of the EAB invasion. This program has now documented successful establishment of the egg parasitoid O. agrili and the two larval parasitoid T. planipennisi and S. galinae in EAB-infested forests at most release sites in the northern USA, in areas where surveys to document parasitoid establishment have been conducted. While the role of O. agrili in suppressing EAB population growth requires continued evaluation, the larval parasitoids T. planipennisi (from China) and S. galinae (from the Russian Far East) have become the dominant biotic factors suppressing EAB population growth rates and significantly reducing EAB densities in the aftermath forests in Michigan and several northeastern states, where these parasitoids were released between 2007 and 2017 (Duan et al. 2015b, 2017; 2021; Margulies et al. 2017; Kashian et al. 2018). EAB densities at these biocontrol study sites are now sufficiently low (< 10 larvae per m2 phloem area) to allow the surviving trees and saplings to recover and grow to canopy trees, reaching the overstory (Duan et al. 2015b, 2017, 2021). We expect that the suppression of EAB densities is likely to expand geographically as established populations of O. agrili, T. planipennisi, and S. galinae increase and spread to new areas and parasitoids are released as part of the ongoing biocontrol release effort. However, tree regrowth and regeneration are very slow processes, normally taking decades. Even after EAB densities are successfully reduced and the frequency of EAB outbreaks moderated by the introduced agents, long-term monitoring studies will be needed to fully assess the contribution of EAB biocontrol to the recovery and regeneration of North American ash in the post-EAB invasion forests. We recommend expanding the current EAB biocontrol research to (1) quantify the long-term impact of EAB biocontrol on ash community and forest recovery in the aftermath of EAB as ash trees grow to canopy size, (2) determine parasitoid establishment in EAB populations in warmer regions (southern United States), (3) explore different regions of Asia for EAB natural enemies adapted to climate zones similar to those in the southern and western USA where EAB is now invading, and (4) evaluate the potential of area-wide EAB control with integration of host plant resistance and/or selective insecticide use (e.g., Davidson and Rieske 2016; Koch et al. 2021).