Anoplophora glabripennis
(Asian longhorned beetle)

Widespread planting of susceptible poplars within the native zone led quite rapidly to the build-up and spread of A. glabripennis, which was previously of no particular importance there. In China, the species has only been a problem since the 1980s, when the first significant outbreak was discovered in Ningxia province (Pan HongYang, 2005). The Asian longhorned beetle A. glabripennis was first intercepted outside its native range in 1992 in North America, where an established population was first observed in 1996 (Haack et al., 1996) in the Brooklyn area of New York (USA). The first European invasive population was detected in Austria in 2001. Since then populations have been detected in Canada as well as in several European countries (Javal et al., 2019a; EPPO, 2020).

A. glabripennis is indigenous to East Asia. In China, the species is considered as native throughout the country. However, its prevalence and range has increased as a result of widespread planting of susceptible poplar hybrids (Pan HongYang, 2005). The pest has therefore a remarkably broad distribution in China, and can be found in most of the country at varying abundances (Yan, 1985; Li and Wu, 1993). The species has also been recorded throughout Korea (Lingafelter and Hoebeke, 2002) but is present at a moderately low density in riparian areas (Williams et al., 2004a). Despite some museum records which suggested that the species could have been native to Japan, it was established that A. glabripennis was not part of the Japanese endemic fauna (Lingafelter and Hoebeke, 2002). A. glabripennis has never been recorded in Taiwan (EPPO, 2020) despite extensive surveillance.

A. glabripennis has been detected in North America. In the USA, the first established population was discovered in New York City in 1996 (Haack et al., 1996). Several other populations have then been detected in the north-eastern part of the country as well as in Canada (Carter et al., 2010, EPPO, 2020). In Europe, established populations have been detected since 2001 (Hérard et al., 2006) and the species is now present in multiple other European countries (Javal et al., 2019a; EPPO, 2020).

Records of A. glabripennis in Maryland, Pennsylvania and Georgia (Nowak et al., 2001) published in previous versions of the Compendium are invalid as Nowak et al. (2001) does not mention detection of the pest in these states. For information on the states where ALB is considered present and under quarantine, see: https://www.aphis.usda.gov/aphis/resources/pests-diseases/asian-longhorned-beetle/Quarantines

The invasion process of the Asian longhorned beetle may have already started in its native zone, since its distribution has been altered by anthropic activities in Asia, as suggested in China (Javal et al., 2019b) and in the Korean peninsula (Lee et al., 2020). In Asia, niche modelling shows that the species is likely to settle in eastern and central China, North and South Korea, and Japan, and has a lower probability to establish in south-east China and east India (Peterson et al., 2004).

The steady increase in international trade from Asia since the early 1980s (Normile, 2004) has certainly increased the likelihood of unintentional introductions of the species. Once introduced, transportation corridors and abundance of preferred hosts are predictors of the ability of the species to spread (Shatz et al., 2013). In addition, the comparison between the climatic niche occupied by the species in its native zone and in its zone of introduction reveals only a negligible shift and expansion of niche (Hill et al., 2017), indicating an absence of changes physical and biological demands of the species during the invasion process.

Attention was drawn to A. glabripennis by its introduction into the USA in 1996 (Haack et al., 1996), where a major eradication programme is underway, and strong measures have been taken to reduce the risk of further introduction with wood packing from China. Niche modelling shows that most of the eastern USA has a high risk of Asian longhorned beetle establishment (Peterson and Vieglais, 2001; Peterson et al., 2004).

As forecasted by the detailed pest risk analysis done by MacLeod et al. (2002), the species was then introduced in Europe in 2001 (EPPO, 2020). Indeed, using the climate-matching system CLIMEX (Skarratt et al., 1995) and the EPPO Standard for pest risk assessment (EPPO, 1997), areas in southern Europe had previously been identified as those where the pest was most likely to establish and cause economic damage. More recent studies of climatic and environmental suitability show that the species might be able to spread in most of Europe, except for the most northern areas (EFSA, 2019). The invasion in Europe is most likely due to several introductions, both from the native zone, but also from already invaded areas such as the USA, highlighting the fact that international trade from all areas where the species is recorded should be the focus of management measures (Javal et al., 2019b).

The species is very polyphagous and is found on a wide range of tree species (MacLeod et al., 2002; Haack et al., 2006; Hu et al., 2009; Sjöman et al., 2014; Meng et al., 2015). The genera most often described as preferential hosts for A. glabripennis include poplar (Populus spp.), willow (Salix spp.), maple (Acer spp.), birch (Betula spp.), elm (Ulmus spp.) and plane tree (Platanus spp.) (Sjöman et al., 2014). In China, A. glabripennis has mainly been recorded on poplar, willow, elm and maple trees (Pan HongYang, 2005) and the major hosts are species and hybrids of section Aegeiros of the genus Populus: P. nigra, P. deltoides, P. x canadensis and the Chinese hybrid P. dakhuanensis. Some poplars of the other sections of the genus (Alba and Tacamahaca) are also attacked, but are only slightly susceptible (Li and Wu, 1993). The insect has never been found on conifers, nor apparently on such important forest genera as Fagus and Quercus. This great diversity of hosts is explained by the peculiarities of the microbiome of its digestive system which allow it both to increase the rate of degradation of certain components of the wood, but also to access essential nutrients that the insect cannot synthesize on its own or metabolize from its diet (Scully et al., 2014). The great polyphagia of A. glabripennis is also linked to regulation mechanisms of genes involved in particular in the digestion and assimilation of nutrients to compensate for development in wood that is poor in nutrients (Mason et al., 2016). The amplification and functional divergence of genes associated with specialization of food on plants, including genes inherited from fungi or bacteria by horizontal transfer, have also contributed to the expansion of A. glabripennis’ metabolic repertoire (McKenna et al., 2016).

Research has been evaluating which tree species are most at risk from larval feeding in the invaded range. Bancroft et al. (2002), rearing larvae in freshly cut logs, ranked eight tree species in the following order for larval weight gain (from largest to smallest): Ulmus chinensis [U. parvifolia], Acer platanoides, Ulmus americana, Gleditsia triacanthos, Acer saccharum, Quercus rubra, Fraxinus americana and Fraxinus pennsylvanica. Smith et al. (2002) found that adult survival and reproductive capacity were higher on A. platanoides than on Acer rubrum and Salix nigra. Ludwig et al. (2002) investigated oviposition under caged conditions, and insertion of first-instar larvae into potted trees as experimental methods for determining host potential; they showed that eggs are laid on, and larvae develop in, species which are not yet known to be hosts (e.g. Q. rubra). Morewood et al. (2003) showed in laboratory experiments that the number of oviposition sites and living larvae was significantly higher in sugar maple (A. saccharum) than on red maple (A. rubrum), green ash (F. pennsylvanica) or red oak (Q. rubra).

Adult specimens can be found between April and December (Haack et al., 2010), but the most active period for adult activity is late June to early July (Li and Wu, 1993) and oviposition most often occurs in early summer. Females form an oviposition well with their mandibles in tree bark, usually at the base of the lowest branch crown (Li and Wu, 1993; Haack et al., 2006). About 45-62 eggs are laid one by one by females (Wong and Mong, 1986; Keena, 2002) between the bark and the cambium. Eggs develop faster at higher temperatures, and take between 54.4 and 13.3 days to hatch at temperatures ranging from 15 to 30°C (Keena, 2006). While developing, the larva first digs galleries under the bark before digging deeper inside the trunk, where it eventually forms a pupal compartment to spend the 12 to 50 days necessary for its metamorphosis (Keena and Moore, 2010). Another 4-7 days are needed for sclerotization to be complete, and 4-5 days for the adult insect to burrow the trunk to the surface and extricate itself from the trunk (Sánchez and Keena, 2013). The emergence of adults lasts throughout the summer, with peak emergence most often between May and June, varying according to region and climatic conditions (Haack et al., 2010). Short-range pheromones allow partner recognition (Zhang et al., 2002; 2003a). The longevity and fecundity of insects are conditioned by the larval host and temperature conditions, but adults usually live for about a month (Li and Wu, 1993; Keena 2002, 2006; Morewood et al., 2003). The life-cycle takes place in 1 to 2 years, depending on the environmental conditions: the individuals spend the winter in larval form, and it is necessary that the larvae have reached a minimum mass at the beginning of winter in order to be able to induce pupation the following spring (Keena and Moore, 2010). In China, the number of annual generations varies with climate and latitude. The further north A. glabripennis is found, the longer it takes for a generation to develop. In eastern China, a generation may take 1 or 2 years to develop, whereas in northern China (Neimenggu), a single generation takes 2 years to develop. Thus, there can be one or two overlapping generations per year, depending upon the climate and feeding conditions (Pan HongYang, 2005). The adults usually remain on the tree from which they emerged, or fly short distances to nearby trees, and feed there on leaves, petioles and young bark before mating and laying eggs (Smith et al., 2001, 2004; Pan HongYang, 2005).

In the native range, most populations are found in urban environments, mainly in range trees and in parks, as well as in monocultures (Smith et al., 2009) but in some cases they can also be observed on the edge of the forest (Williams et al., 2004a). In Europe, no outbreaks have been detected outside urban areas (EPPO, 2020). The situation is similar in North America, with the exception of a population detected in 2008 in a peri-urban forest in Massachusetts (Dodds and Orwig, 2011). Given the current distribution of the species, its biology, and its regular interceptions around the world, it is possible that A. glabripennis could become established in much of Europe (MacLeod et al., 2002), but also in North America and Asia (Hu et al., 2009). On the basis of this information, the species has been classified as a quarantine pest in both Europe (EPPO, 2020) and North America (USDA, 1998).

Luo et al. (2003) briefly review possibilities for biological control of the pest. Multiple control strategies are still under study, but do not seem conclusive at the moment (Pan HongYang, 2005; Dubois et al., 2008; Haack et al., 2010; Ugine et al., 2013; Brabbs et al., 2015) often due to their toxicity or their low specificity of action. Indeed, A. glabripennis has few natural enemies because the majority of its development cycle takes place hidden inside trees, making its management by biological control methods difficult (Pan HongYang, 2005). For example, entomopathogenic bacteria have been identified as a potential means of controlling insects (Pan HongYang, 2005; Brabbs et al., 2015) but the tests carried out to date in the laboratory have not revealed any significant effect on the different stages of development of individuals (D'Amico et al., 2004). Some entomopathogenic fungi of the genera Beauveria and Metarhizium are also known for their virulence, and have the advantage of being able to form epidemics within populations (Shimazu et al., 2002; Zhang et al., 2003b; Dubois et al., 2008; Hu et al., 2009; Ugine et al., 2014; Brabbs et al., 2015) and some strains are even registered as biocontrol agents in several regions of the world (Dubois et al., 2008). Several nematode species also attack A. glabripennis, mainly belonging to the genus Steinernema (Fallon et al., 2004) and can trigger high level of mortality (Liu et al., 1992). However, their use as a biological control agent is made tricky by the difficulty of maintaining certain species under laboratory conditions, but also by their sensitivity to the method of release into the environment which largely conditions their survival rate (Brabbs et al., 2015). The option of parasitoids was also explored. Among the known species of Anoplophora parasitoids, larvae of the beetle Dastarcus helophoroides are ectoparasites of A. glabripennis larvae (Pan HongYang, 2005; Hu et al., 2009; Haack et al., 2010; Brabbs et al., 2015). In Europe, native parasitoids attacking the larvae of A. glabripennis have been identified, but their dissemination in urban areas is made impossible for public health reasons (Brabbs et al., 2015). Finally, A. glabripennis has few predators. No study shows any predation on A. glabripennis by any other insect species (Haack et al., 2010). Some woodpecker species, on the other hand, can feed on the larvae. The grey-headed woodpecker (Picus canus) and the great spotted woodpecker (Dendrocopos major) for instance, are both found in both Asia and Europe, and are effective predators of A. glabripennis. Their nesting is therefore encouraged as much as possible (Pan HongYang, 2005).

Based on random amplified polymorphic DNA (RAPD) fragments, Kethidi et al. (2003) describe DNA markers for the molecular identification of all development stages, and frass, of A. glabripennis, as distinct from related species. Other RAPD markers have also been used for intraspecific differentiation of populations (An et al., 2004). Molecular barcoding, based on mitochondrial DNA, provides a powerful tool to identify specimens, including immatures, to the species level (Javal et al., 2019a). Microsatellite markers have also been developed and used to visualize population structure and to decipher invasion routes (Carter et al., 2008, 2009, 2010; Javal et al., 2019b). A rapid diagnostic protocol based on loop-mediated isothermal amplification (LAMP) has been developed (Rizzo et al., 2020).

A. glabripennis is very similar to A. freyi from south-western China (Lingafelter and Hoebeke, 2002; Wu WeiWen and Chen Bin, 2003) and to A. coeruleoantennata (Lingafelter and Hoebeke, 2002).

In its invaded zone, A. glabripennis can be confused with pine sawyer species (Monochamus spp.). However, Monochamus species are usually smaller, usually do not show the same elytral patterns, and have a different phenology. The cotton borer is another similar species; however, it is characterized by black spots on light yellow elytra. In addition, its North-American distribution does not currently overlap with that of A. glabripennis (Meng et al., 2015). The greatest risk of confusion is with another invasive species, Anoplophora chinensis (citrus longhorned beetle, synonymous with A. malasiaca) (Haack et al., 2010; Pennacchio et al., 2012). Adults of A. chinensis differ from A. glabripennis in the presence of tubers at the base of their elytra, and the larvae of the two species show patterns of different size and shape on the disc of their pronotum (Lingafelter and Hoebeke, 2002; Pennacchio et al., 2012). A. chinensis has also been categorized as a quarantine pest by European countries (EPPO, 1997).

Due to the variable regulations around (de)registration of pesticides, your national list of registered pesticides or relevant authority should be consulted to determine which products are legally allowed for use in your country when considering chemical control. Pesticides should always be used in a lawful manner, consistent with the product's label.

In China, control measures include the direct application of insecticides (Chen et al., 1990; Liang et al., 1997), trap trees combined with insecticide treatments (Sun et al., 1990) or the use of insect-pathogenic nematodes (providing up to 94% mortality; Liu et al., 1992). As certain poplar hybrids are relatively resistant (Qin et al., 1996), the planting of such hybrids is now preferred, and the use of very susceptible hybrids is avoided. Control strategies in China have been reviewed by Luo et al. (2003) and Pan HongYang (2005).

The case of A. glabripennis has been the main stimulus for the development by the FAO Interim Commission on Phytosanitary Measures of an international standard 'Guidelines for regulating wood packaging material in international trade' (ICPM, 2002). In the USA and in Europe, strong measures have been taken for wood packing materials (packing cases and dunnage) from China. Such packaging should be treated by methods recognized to have adequate efficacy against all wood pests. These currently include heat treatment (to an internal temperature of 56°C for 30 min). Once treated, packing wood is unlikely to be re-infested, so such wood (especially crates and pallets) can continue to be used in trade. An internationally recognized mark is stamped on the treated wood. The international standard was approved in 2002 and is now progressively being implemented worldwide, in order to prevent introductions.

In the invaded zones, control measures aim to contain and eradicate the outbreaks in urban areas (Lance, 2003; Haack et al., 2010). Unger (2003) gives an example of the measures needed to exclude the pest from Germany. Detection of outbreaks is difficult and is often incidental. Indeed, the cryptic lifestyle and tendency of the beetle to lay small numbers of eggs on several trees combine to make it difficult to define the limits of the outbreak and thus to eradicate the beetle without destroying large numbers of trees. Once the presence of the species is confirmed, a security perimeter is defined around the infested trees, and these trees are immediately felled and then crushed or incinerated on site. Host species in the vicinity of the outbreak are usually controlled visually. Felling of infested trees is currently the only means of control implemented systematically in Europe, Canada and the USA (Hérard et al., 2006; Meng et al., 2015; Turgeon et al., 2015). This method has proven to be effective in relatively recent, small-scale outbreaks. Inspection of susceptible trees is, however, often not sufficient to detect the entire infestation. Indeed, as the most obvious sign of the presence of the species is the emergence hole, the infestation is very often detected when the adults have already emerged from the tree and have potentially dispersed further. In addition to the visual inspection, dogs have been trained to detect the presence of A. glabripennis (Hoyer-Tomiczek et al., 2016). These dogs are capable of smelling the insect in trees, and are also used to inspect suspicious shipments at potential points of entry.

The difficulty in detecting infestations explains, for example, why new infested trees were detected in 2013 in the Toronto area (10 years after the first discovery of the species in the region) a few months after the species was considered as eradicated from the monitoring area. Growth rings from felled infested trees showed that the individuals had been present for several years, suggesting that this second finding was probably a satellite population of the original population observed (Turgeon et al., 2015).

09/11/20 Updated by:

Marion Javal, Stellenbosch University, South Africa