TY - JOUR
AU - Cipollini,, D
AB - Abstract Emerald ash borer (EAB, Agrilus planipennis Fairmaire [Coleoptera: Buprestidae]) is a wood boring beetle that is an invasive pest of ash trees (Fraxinus spp.) in North America. In 2014, it was reported that EAB had infested white fringetree (Chionanthus virginicus L. [Lamiales: Oleaceae]) in Ohio and was since found to have infested this species across its invasive range. In 2018, we reexamined 166 white fringetrees in Illinois, Indiana, Ohio, and Pennsylvania that had been previously examined for EAB attack in 2015 to determine their fate. We assessed tree health and EAB infestation in each tree, assigned an infestation status of newly, continuously, not reinfested, or never infested, and compared the trees’ current status to their 2015 status. This assessment was done to determine whether their health and infestation status had changed through the EAB invasion wave. We found that attack rates declined: 26% of trees were infested in 2015 whereas only 13% were in 2018, likely coinciding with declining beetle populations in the area. Overall tree health improved for trees that were not reinfested by EAB after a record of attack in 2015, suggesting that they can survive and recover from EAB attack. Conversely, health declined for newly and continuously infested trees, indicating that they became stressed from EAB attack. Although the majority of the trees survived the invasion wave, several were removed from various sites due to EAB attack suggesting that white fringetree varies in its resistance and tolerance to attack. As beetle populations continue to expand geographically, infestation rates will likely increase and health of white fringetrees will decrease with the EAB attack wave, especially as EAB reaches denser populations of fringetrees. host shift, ecological fitting, invasive pest, infestation, tolerance The fate of a host tree in the face of an invasive insect pest is dependent on several biotic and abiotic factors, including the degree of resistance it expresses to the pest, the degree to which the host is being attacked, and environmental stressors. Several wood-boring beetles in the family Buprestidae are important invasive pests (Haack et al. 2002, Haack 2006, Coleman and Seybold 2008, Flint et al. 2013). These beetles can weaken and contribute to host mortality by girdling trees as larvae feed on vascular tissue. Buprestids tend to attack trees that have been compromised by pathogens, cankers, or environmental stress, such as drought (Goheen and Hansen 1993), but in some instances, can attack, weaken, and kill even healthy trees (Cappaert et al. 2005). Larger, stressed trees are known to be highly susceptible to wood and bark beetle attack and therefore are more likely to be killed than smaller (and younger), healthy trees. The gold spotted oak borer (Agrilus coxalis Waterhouse), the red oak borer (Enaphalodes rufulus Haldeman), and the mountain pine beetle (Dendroctonus ponderosae Hopkins) preferentially attack large, stressed and densely populated oak trees and pine trees, respectively, which leads to high mortality among these size classes (Goheen and Hansen 1993, Fierke et al. 2005a, Coleman et al. 2011). Beetle densities are also directly related to the degree to which a host tree may get attacked. For example, when mountain pine beetle populations are low, only large, stressed pine trees get attacked, but as beetle populations increase, attack rates increase on all size classes (Goheen and Hansen 1993). Once host tree populations are depleted from beetle attack, pressure on surviving individuals is lowered as beetle populations decline in a local area. Host range expansions can be achieved via ecological fitting when an organism uses inherent traits or characteristics to find and successfully utilize a novel host (Janzen 1985, Agosta 2006). A number of host range expansions have been documented among Buprestid beetles. Two lined chestnut borer (Agrilus bilineatus Weber) exemplifies this phenomenon because after its primary host, American chestnut, was decimated by chestnut blight, these beetles have primarily utilized stressed oak trees as a host (Dunn et al. 1986). Emerald ash borer (Agrilus planipennis Fairmaire) has also undergone a host range expansion. This invasive, Asian wood boring beetle is considered a specialist on ash trees (Fraxinus spp.) using Manchurian ash (Fraxinus mandschurica Rupr.) and a few other species as its primary ancestral hosts. Upon its arrival to North America in the 1990s (Siegert et al. 2014), EAB began to attack and decimate North American ash tree populations throughout the Eastern United States and southern Canada (Haack et al. 2002). In 2014, it was discovered that EAB had been attacking another novel host in North America: white fringetree (Chionanthus virginicus L.; Cipollini 2015), the first nonash host confirmed for this beetle. White fringetrees can be classified as trees or shrubs depending on the growth form of the individual; they grow up to 11 meters in height with an often multi-stemmed growth habit (USDA NRCS 2019). White fringetree is in the same family and closely related to ash trees (Oleaceae; Wallander and Albert 2000) and are apparently physically and chemically similar enough for EAB to utilize it as a host via ecological fitting (Cipollini and Peterson 2018). The native range of white fringetree extends as far north as New York, as far south as Florida and as far west as Texas; thus, it grows alongside ash trees in its native habitats throughout North America, and they are planted as ornamentals within and outside of their native range in many of the same areas where ash trees grow and are planted (USDA NRCS 2019). White fringetree is mostly seed propagated and there are few recognized cultivars currently for sale, partly because of the difficulty in propagating them via cuttings (Russ 2010). While the fate of North American ash trees in the face of EAB has been examined in a number of studies (Tanis and McCullough 2012, Spei and Kashian 2017), the fate of white fringetrees has only begun to be examined. When assessed in 2015, 26% of 178 white fringetrees growing in ornamental landscapes in Ohio, Illinois, Indiana, and Pennsylvania were found to be attacked by EAB, with most showing signs of canopy decline and a few trees dead due to EAB attack (Peterson and Cipollini 2017). At nearly all of the planted sites examined, at least some white fringetrees were found to be attacked, and trees had a higher probability of being attacked if they were larger, growing in denser populations, and had epicormic sprouting and canopy dieback (both symptoms of attack and indicators of stress that may promote attack; Peterson and Cipollini 2017). A dendrochronological study indicated that white fringetrees in southwestern Ohio first got attacked at about the same time as ash trees were attacked in the local area (Thiemann et al. 2016). Unfortunately, each of these studies was limited by being retrospective with no prior knowledge of the condition of trees assessed in each study and the exact timing of attack on them. In this study, we reassessed the ornamental white fringetrees that were assessed in 2015 (Peterson and Cipollini 2017) to determine their fate through the invasion wave of EAB over the following 3 yr. Their fate was determined by examining changes in infestation status as well as changes in the degree of canopy dieback and epicormic sprouting that could be observed since they were first assessed. We also examined how well factors such as tree size, canopy condition, and infestation status in 2015 could predict infestation status in 2018. Since beetle densities have largely declined in the areas encompassing our inventory of trees, we predicted that the majority of the trees would not be infested upon the 2018 reassessment. In turn, we predicted that tree health would generally stay the same or improve in these trees but would decline in those continually or newly infested. We also predicted that current year attack would be predicted by both past attack and tree health. Finally, we sought to expand the current map of white fringetree with signs of EAB attack in eastern North America. We predicted that the range of sites with evidence of EAB attack on white fringetree would continue to expand as the beetle has expanded its range. Materials and Methods White Fringetree Sites and Reassessment We reexamined the same ornamental trees at the same sites used by Peterson and Cipollini (2017) in Illinois, Indiana, Ohio, and Pennsylvania (Table 1, Fig. 1). Trees were located using our own tree inventory records along with the help of property managers. The majority of the fringetrees were open grown in well-manicured gardens, cemeteries, and arboretums; some showed evidence of periodic pruning and were often planted in mulch beds or mowed grass lawns, but none of them were treated with insecticides, watered or fertilized as affirmed via personal communication with land managers. We saw no evidence that any trees in our study had been grafted, a technique that has generally not been used successfully for this species (University of Arkansas Cooperative Extension Factsheet 2006). Otherwise we had very limited information on the source of the trees planted in most of our sites. We reassessed a total of 166 of the original 178 trees due to the removal of 12 dead or unhealthy fringetrees from various sites since our initial survey. Table 1. Study sites in Illinois, Indiana, Ohio, and Pennsylvania with GPS coordinates, total number of white fringetrees (Chionanthus virginicus) present in each site, and number of trees for each infestation status Site . State . GPS Coordinates . Number of trees . Infestation status . . . . . Newly . Continuously . Not reinfested . Never . Chicago Botanic Gardens IL 42.1491°N, 87.7894°W 25 0 0 1 24 Cox Arboretum OH 39.6554°N, 84.2243°W 18 4 3 1 10 Hershey Gardens PA 40.2859°N, 76.6502°W 5 1 0 1 3 Indianapolis Museum of Art: Newfields IN 39.8260°N, 86.1857°W 10 0 0 3 7 Morton Arboretum IL 41.8164°N, 88.0549°W 14 1 0 5 8 Purdue University IN 40.4237°N, 86.9212°W 20 1 4 1 14 Ferncliff Cemetery OH 39.9242°N, 83.8088°W 4 0 0 3 1 Spring Grove Cemetery OH 39.1743°N, 84.5250°W 36 0 1 11 24 Wilmington College OH 39.4448°N, 83.8182°W 2 0 0 1 1 Woodland Cemetery OH 39.7442°N, 84.1728°W 2 1 0 1 0 Wright Brother’s Memorial OH 39.4738°N, 84.0517°W 1 0 1 0 0 Wright Memorial Public Library OH 39.7156°N, 84.1712°W 3 0 0 0 3 Wright-Patterson Air Force Base OH 39.8137°N, 84.0537°W 8 0 0 1 7 Yellow Springs OH 39.8064°N, 83.8869°W 18 2 3 4 9 Total 166 10 12 33 111 Site . State . GPS Coordinates . Number of trees . Infestation status . . . . . Newly . Continuously . Not reinfested . Never . Chicago Botanic Gardens IL 42.1491°N, 87.7894°W 25 0 0 1 24 Cox Arboretum OH 39.6554°N, 84.2243°W 18 4 3 1 10 Hershey Gardens PA 40.2859°N, 76.6502°W 5 1 0 1 3 Indianapolis Museum of Art: Newfields IN 39.8260°N, 86.1857°W 10 0 0 3 7 Morton Arboretum IL 41.8164°N, 88.0549°W 14 1 0 5 8 Purdue University IN 40.4237°N, 86.9212°W 20 1 4 1 14 Ferncliff Cemetery OH 39.9242°N, 83.8088°W 4 0 0 3 1 Spring Grove Cemetery OH 39.1743°N, 84.5250°W 36 0 1 11 24 Wilmington College OH 39.4448°N, 83.8182°W 2 0 0 1 1 Woodland Cemetery OH 39.7442°N, 84.1728°W 2 1 0 1 0 Wright Brother’s Memorial OH 39.4738°N, 84.0517°W 1 0 1 0 0 Wright Memorial Public Library OH 39.7156°N, 84.1712°W 3 0 0 0 3 Wright-Patterson Air Force Base OH 39.8137°N, 84.0537°W 8 0 0 1 7 Yellow Springs OH 39.8064°N, 83.8869°W 18 2 3 4 9 Total 166 10 12 33 111 Open in new tab Table 1. Study sites in Illinois, Indiana, Ohio, and Pennsylvania with GPS coordinates, total number of white fringetrees (Chionanthus virginicus) present in each site, and number of trees for each infestation status Site . State . GPS Coordinates . Number of trees . Infestation status . . . . . Newly . Continuously . Not reinfested . Never . Chicago Botanic Gardens IL 42.1491°N, 87.7894°W 25 0 0 1 24 Cox Arboretum OH 39.6554°N, 84.2243°W 18 4 3 1 10 Hershey Gardens PA 40.2859°N, 76.6502°W 5 1 0 1 3 Indianapolis Museum of Art: Newfields IN 39.8260°N, 86.1857°W 10 0 0 3 7 Morton Arboretum IL 41.8164°N, 88.0549°W 14 1 0 5 8 Purdue University IN 40.4237°N, 86.9212°W 20 1 4 1 14 Ferncliff Cemetery OH 39.9242°N, 83.8088°W 4 0 0 3 1 Spring Grove Cemetery OH 39.1743°N, 84.5250°W 36 0 1 11 24 Wilmington College OH 39.4448°N, 83.8182°W 2 0 0 1 1 Woodland Cemetery OH 39.7442°N, 84.1728°W 2 1 0 1 0 Wright Brother’s Memorial OH 39.4738°N, 84.0517°W 1 0 1 0 0 Wright Memorial Public Library OH 39.7156°N, 84.1712°W 3 0 0 0 3 Wright-Patterson Air Force Base OH 39.8137°N, 84.0537°W 8 0 0 1 7 Yellow Springs OH 39.8064°N, 83.8869°W 18 2 3 4 9 Total 166 10 12 33 111 Site . State . GPS Coordinates . Number of trees . Infestation status . . . . . Newly . Continuously . Not reinfested . Never . Chicago Botanic Gardens IL 42.1491°N, 87.7894°W 25 0 0 1 24 Cox Arboretum OH 39.6554°N, 84.2243°W 18 4 3 1 10 Hershey Gardens PA 40.2859°N, 76.6502°W 5 1 0 1 3 Indianapolis Museum of Art: Newfields IN 39.8260°N, 86.1857°W 10 0 0 3 7 Morton Arboretum IL 41.8164°N, 88.0549°W 14 1 0 5 8 Purdue University IN 40.4237°N, 86.9212°W 20 1 4 1 14 Ferncliff Cemetery OH 39.9242°N, 83.8088°W 4 0 0 3 1 Spring Grove Cemetery OH 39.1743°N, 84.5250°W 36 0 1 11 24 Wilmington College OH 39.4448°N, 83.8182°W 2 0 0 1 1 Woodland Cemetery OH 39.7442°N, 84.1728°W 2 1 0 1 0 Wright Brother’s Memorial OH 39.4738°N, 84.0517°W 1 0 1 0 0 Wright Memorial Public Library OH 39.7156°N, 84.1712°W 3 0 0 0 3 Wright-Patterson Air Force Base OH 39.8137°N, 84.0537°W 8 0 0 1 7 Yellow Springs OH 39.8064°N, 83.8869°W 18 2 3 4 9 Total 166 10 12 33 111 Open in new tab Fig. 1. Open in new tabDownload slide Locations where emerald ash borer (Agrilus planipennis) has been found attacking ornamental (circle) and wild (triangle) white fringetree (Chionanthus virginicus) populations in the eastern and midwestern United States. Fig. 1. Open in new tabDownload slide Locations where emerald ash borer (Agrilus planipennis) has been found attacking ornamental (circle) and wild (triangle) white fringetree (Chionanthus virginicus) populations in the eastern and midwestern United States. Each fringetree was reexamined once in the summer (June, July, and August) of 2018 for canopy dieback, epicormic sprouting, and signs and symptoms of past or current emerald ash borer attack (Fig. 1). Every ornamental white fringetree in this study was examined via personal observations for EAB infestation, as we have done previously (Peterson and Cipollini 2017). The signs of EAB infestation include the presence of a D-shaped exit hole on stems or branches, feeding larvae or serpentine larval galleries beneath the bark containing larval frass. Symptoms of EAB attack on trees include circumferential swellings of stems or branches over feeding galleries, epicormic sprouting, and canopy dieback. If no external visible signs of attack were obvious, but there was suspicious bark swelling or other deformities, that particular area of the stem or branch was carefully debarked using a wood chisel to confirm gallery formation (Thiemann et al. 2016). Determining whether attack was new or old was based on inventory records from Peterson and Cipollini (2017) and on how deep the gallery was under the vascular cambium. If the gallery was near the newest layer of phloem tissue, more brown in color, had visible frass or feeding larvae, then it was considered new. If the gallery was under several layers of vascular tissue, appeared black in color, and had no new frass or larvae, then it was considered old. A tree could have both new and old galleries; thus, we removed several layers of secondary xylem and phloem around suspicious areas to reveal galleries. Infestation status of each tree was categorized as newly infested, continuously infested, not reinfested, or never infested. Newly infested trees were trees that showed no signs of infestation by EAB in 2015 but were attacked at some point since then. Continuously infested trees were classified as infested in 2015 and again in 2018, whereas not reinfested trees were classified as infested in 2015, but with no signs of new infestation since then. Last, white fringetrees with no signs of infestation during either assessment were considered to be never infested. Canopies were visually assessed for dieback by one of the authors, which was quantified as a percentage of the canopy missing leaves. Each tree was placed into one of five categories corresponding to their percent dieback: 1 = 0% dieback; 2 = 25% dieback; 3 = 50% dieback; 4 = 75% dieback; and 5 = 100% dieback, based on the Gould et al. (2015) ash tree rating system. The presence or absence of epicormic sprouts, which grow from dormant buds beneath the bark of the tree that are activated and produce new branches when the tree is damaged or stressed, was recorded for each tree. Trees that were pruned for aesthetic reasons were documented in the event that the tree had lost epicormic sprouts recorded in 2015 due to pruning because trees typically do not lose epicormic sprouts that quickly, if at all. We did not re-measure the diameter of the largest stem (a metric recorded in the initial survey) due to difficulty in detecting the precise location of the previous measurement, and because these trees expand in diameter slowly. Statistical Analysis In our statistical analyses, we used data from both 2015 and 2018 to examine factors affecting the condition of trees in 2018, but otherwise analyzed these data in largely the same way as in Peterson and Cipollini (2017). Site was not included in our models in the original study or in this analysis due to vastly unequal replication of trees within sites. An ANOVA using PROC GLM and Tukey HSD Post Hoc test was used to examine variation in the diameter of the largest stem (measured in 2015), canopy dieback, and percent change in canopy dieback between 2015 and 2018 among infestation statuses, as determined in 2018. A χ 2-test with PROC FREQ was used to compare the frequency of epicormic sprouting among infestation status categories as observed on trees in 2015 and 2018. We used binary regression with PROC LOGISTIC to determine relationships between current infestation status and the presence of epicormic sprouts (measured in 2015 and 2018), basal diameter of the largest stem (measured in 2015), crown dieback (measured in 2015 and 2018) and previous infestation status (measured in 2015). Stepwise model selection was used to select the best model with AIC. Data were analyzed using SAS (SAS Studio Institute Inc., Cary, NC, 2018). We published an initial distribution map of white fringetrees showing signs of attack by EAB in Peterson and Cipollini (2017), based on our observations through 2016. We sought to expand upon this map by adding newly discovered trees with signs of attack. This was not a systematic survey, and trees were not fully assessed as in our main study. In most cases, specific locations were visited to examine planted white fringetrees that were previously unknown to us and in some cases, trees were discovered opportunistically and assessed for signs of EAB attack. We confirmed any new infestations discovered through personal observations, or in two instances, relied on photographic evidence from colleagues. The GPS coordinates of each previously and newly discovered site with at least one infested white fringetree was documented and mapped using QGIS software. Results Overall, we found that attack rates declined from 2015 to 2018. In 2015, 47 white fringetrees were infested out of 178 (Peterson and Cipollini 2017). Only 12 of the fringetrees were reinfested in 2018 and 10 were newly infested of 166 trees remaining in the study (Table 1). The majority of the trees were never infested (111) while 33 of the previously infested trees were not reinfested (Table 1). The 12 trees that were not reassessed in 2018 had been removed due to poor health, and in most cases, due to severe EAB infestation. One of the trees was removed in Yellow Springs, OH, due to declining canopy health due to infestation by EAB. Two of the trees on Wright Patterson Airforce Base in Fairborn, OH, were removed due to declining canopy health, but neither had confirmed EAB attack. Five trees were removed from Morton Arboretum in Lisle, IL, due to declining canopy health: three of these were removed due to EAB infestation, and two trees were removed for unknown reasons. Last, two trees from the Chicago Botanic Gardens in Glencoe, IL, were removed due to EAB infestation. Average stem diameter varied among trees with different infestation statuses. On average, continuously attacked and not reinfested trees had the largest stem diameter and were >1.5 times larger than never infested trees (F3,161= 8.82, P < 0.001; Table 2). Newly attacked trees were slightly larger than never infested trees, but not significantly different than either continuously or not reinfested trees (Table 2). Table 2. Mean tree diameter, percent canopy dieback in 2015 and 2018, and change in canopy dieback from 2015 to 2018 ± SD of ornamental white fringetrees (Chionanthus virginicus) across sites in IN, IL, OH, and PA. Different lowercase letters indicate significant differences via ANOVA with Tukey’s HSD posthoc test P < 0.05) Variable . Infestation status . . Newly . Continuously . Not reinfested . Never . Avg diam (cm) 2015 6.61 ± 2.08ab 10.07 ± 4.44b 10.76 ± 5.87b 6.44 ± 4.24a Avg % canopy dieback 2015 11.11 ± 18.16ab 15 ± 19.54ab 26.61 ± 23.22a 10.17 ± 19.66b Avg % canopy dieback 2018 27.77 ± 15.02a 27.08 ± 22.51a 16.13 ± 21.92b 12.83 ± 21.68b Change in canopy dieback from 2015 to 2018 -16.67 ± 21.65a -12.08 ± 27.42a 10.48 ± 25.64b -2.65 ± 20.42a Variable . Infestation status . . Newly . Continuously . Not reinfested . Never . Avg diam (cm) 2015 6.61 ± 2.08ab 10.07 ± 4.44b 10.76 ± 5.87b 6.44 ± 4.24a Avg % canopy dieback 2015 11.11 ± 18.16ab 15 ± 19.54ab 26.61 ± 23.22a 10.17 ± 19.66b Avg % canopy dieback 2018 27.77 ± 15.02a 27.08 ± 22.51a 16.13 ± 21.92b 12.83 ± 21.68b Change in canopy dieback from 2015 to 2018 -16.67 ± 21.65a -12.08 ± 27.42a 10.48 ± 25.64b -2.65 ± 20.42a Open in new tab Table 2. Mean tree diameter, percent canopy dieback in 2015 and 2018, and change in canopy dieback from 2015 to 2018 ± SD of ornamental white fringetrees (Chionanthus virginicus) across sites in IN, IL, OH, and PA. Different lowercase letters indicate significant differences via ANOVA with Tukey’s HSD posthoc test P < 0.05) Variable . Infestation status . . Newly . Continuously . Not reinfested . Never . Avg diam (cm) 2015 6.61 ± 2.08ab 10.07 ± 4.44b 10.76 ± 5.87b 6.44 ± 4.24a Avg % canopy dieback 2015 11.11 ± 18.16ab 15 ± 19.54ab 26.61 ± 23.22a 10.17 ± 19.66b Avg % canopy dieback 2018 27.77 ± 15.02a 27.08 ± 22.51a 16.13 ± 21.92b 12.83 ± 21.68b Change in canopy dieback from 2015 to 2018 -16.67 ± 21.65a -12.08 ± 27.42a 10.48 ± 25.64b -2.65 ± 20.42a Variable . Infestation status . . Newly . Continuously . Not reinfested . Never . Avg diam (cm) 2015 6.61 ± 2.08ab 10.07 ± 4.44b 10.76 ± 5.87b 6.44 ± 4.24a Avg % canopy dieback 2015 11.11 ± 18.16ab 15 ± 19.54ab 26.61 ± 23.22a 10.17 ± 19.66b Avg % canopy dieback 2018 27.77 ± 15.02a 27.08 ± 22.51a 16.13 ± 21.92b 12.83 ± 21.68b Change in canopy dieback from 2015 to 2018 -16.67 ± 21.65a -12.08 ± 27.42a 10.48 ± 25.64b -2.65 ± 20.42a Open in new tab The average percent canopy dieback in 2015 and 2018 varied significantly among trees with different infestation statuses, despite there being substantial variability in dieback estimates. In 2015, not reinfested trees, on average, had 2.6 times more canopy dieback than never infested trees (F3,161= 5.79, P = 0.0009; Table 2). In 2018, on average, newly and continuously infested trees had 3.8 times more canopy dieback than not reinfested and never infested trees, which did not differ from each other (F3,161= 3.02, P = 0.0313; Table 2). Change in canopy dieback from 2015 to 2018 was significantly different between not reinfested versus newly, continuously, and never infested trees (F3,161=5.39, P = 0.0015; Table 2). Canopy dieback worsened by 60% in newly infested trees followed by continuously (45%) and never (21%) infested trees (Table 2). Canopy dieback improved in not reinfested trees by 65% (Table 2). We found variation in the frequency of epicormic sprouting in both 2015 (χ 2 = 36.84, P < 0.0001) and 2018 (χ 2 = 15.36, P < 0.0001) among infestation statuses (Table 3). Trees that had not been attacked in 2015 but that were classified as newly infested in 2018 showed significant increases in the frequency of epicormic sprouting. Trees that were infested in 2015 and continuously infested in 2018 had significantly more epicormic sprouting present than absent. The number of trees with epicormic sprouting declined by only 2 from 2015 to 2018, which was due to pruning, but presence of epicormic sprouting on trees remained significantly more common than its absence (Table 3). Not reinfested trees had significantly more epicormic sprouting present than absent in 2015, but this ratio declined in 2018. Never infested trees had significantly less epicormic sprouting present than absent in both 2015 and 2018, with only seven trees first acquiring sprouts between 2015 and 2018 (Table 3). Table 3. Number of white fringetrees (Chionanthus virginicus) displaying the presence or absence of epicormic sprouting across United States field sites for each infestation status in 2015 and 2018 . Frequency of epicormic sprouting by infestation status . . Newly . Continuously . Not reinfested . Never . . Present . Absent . Present . Absent . Present . Absent . Present . Absent . 2015 4 5 12 0 23 8 33 80 2018 7 2 10 2 15 16 40 73 . Frequency of epicormic sprouting by infestation status . . Newly . Continuously . Not reinfested . Never . . Present . Absent . Present . Absent . Present . Absent . Present . Absent . 2015 4 5 12 0 23 8 33 80 2018 7 2 10 2 15 16 40 73 Open in new tab Table 3. Number of white fringetrees (Chionanthus virginicus) displaying the presence or absence of epicormic sprouting across United States field sites for each infestation status in 2015 and 2018 . Frequency of epicormic sprouting by infestation status . . Newly . Continuously . Not reinfested . Never . . Present . Absent . Present . Absent . Present . Absent . Present . Absent . 2015 4 5 12 0 23 8 33 80 2018 7 2 10 2 15 16 40 73 . Frequency of epicormic sprouting by infestation status . . Newly . Continuously . Not reinfested . Never . . Present . Absent . Present . Absent . Present . Absent . Present . Absent . 2015 4 5 12 0 23 8 33 80 2018 7 2 10 2 15 16 40 73 Open in new tab Canopy dieback in 2018, the infestation status in 2015, and epicormic sprouting in 2018 were significant predictors of infestation status in 2018 according to the stepwise regression model. For every 25% increase in canopy dieback in 2018, the trees were 9.6 times more likely to be infested (χ = 4.74, P = 0.029). Fringetrees that were infested in 2015 were 3.8 times more likely to be infested in 2018 than those not previously infested (χ = 6.68, P = 0.009). Fringetrees with epicormic sprouting in 2018 were 5.5 times more likely to be infested in 2018 (χ = 7.81, P = 0.005). The other variables examined were not significant predictors of infestation in 2018. Emerald ash borer attack on both ornamental and wild white fringetrees has been confirmed in five new states and at several new sites in states already known to hold attacked trees (Cipollini, personal observation; Peterson, personal observation; Ellison, personal observation; Hoban, personal communication; Fig. 1, Table 1). Newly documented attacks on ornamental white fringetrees were found in gardens and arboretums in Toledo, OH; Cleveland, OH; West Portsmouth OH; Newark, OH; Wooster, OH; Columbus, OH; Ann Arbor, MI; Bloomington, IN; Lexington, KY; Morgantown, WV; and on a private property in NC (Fig. 1). Infested trees in wild populations were found at the Latodami Nature Center in Wexford, PA, a private property near Jackson, OH, the Vinton Furnace State Experimental Forest in Oreton, OH, and on Bear Island in Potomac, MD (Fig. 1). Trees growing in wild populations typically grow at higher densities and appear to be attacked at higher rates than trees planted at low density in ornamental landscapes, but these observations were not quantified in this study. Discussion We re-examined a group of ornamental white fringetrees first examined 3–4 yr prior to this study to determine the fate of this novel host through the invasion wave of emerald ash borer. Based on our findings, it appears that this species will meet a better fate than most ash trees native to eastern North America. Concurring with our previous study (Peterson and Cipollini 2017), we found that the majority of the white fringetrees that we examined had not and were not being attacked by EAB. Moreover, of the trees that were previously attacked, most of them were not re-infested. This implies that beetle densities are decreasing in the areas encompassed by our survey as the peak of the invasion wave has moved away from these areas (http://www.emeraldashborer.info/timeline/by_county/index.html). Of trees that were attacked, tree stress (e.g., canopy dieback, epicormic sprouting, and previous infestation of the tree) was a good predictor (and sign) of current infestation. Accordingly, we should continue to see higher infestation rates on previously attacked or otherwise stressed trees in locations where there is sufficient beetle pressure, as seen in previous studies of EAB and with other Buprestids (Fierke et al. 2005b, Coleman et al. 2011, Tluczek et al. 2011). Previously attacked trees had higher odds of getting reinfested by EAB, but because beetle pressure in the areas encompassed by our survey has generally declined between 2015 and 2018, the number of reinfested trees in our survey was low. Trees that were reinfested or that were infested for the first time between 2015 and 2018 showed signs of declining health, such as increased canopy dieback and epicormic sprouting between 2015 and 2018, whereas trees not reinfested by EAB showed signs of recovery. The canopies of these trees were able to recover due to decreased stress on the tree and inherent wound healing and canopy regeneration. For those fringetrees experiencing new or continued attack, it can be inferred that their health will decline and that they will be even more prone to attack in the future, but should be able to recuperate if released from attack by the beetle. As predicted, our results provide little evidence that white fringetrees are currently acting as a major reservoir for EAB (Cipollini and Peterson 2018). Oviposition and successful usage of white fringetree by EAB seems to transpire primarily when beetle densities and pressure are high and oviposition spillover occurs (Kaplan and Denno 2007). Because we surveyed areas where beetle densities and pressure were no longer high, further investigation in high pressure areas is required to support this hypothesis. Nonetheless, it appears that white fringetrees will not sustain a sufficiently viable population of EAB when growing in the kind of low density ornamental plantings that we surveyed to serve as a reservoir for EAB until ash tree populations rebound, as can be seen in other invasive pests that use alternative hosts until the availability of their preferred host increases (Lambert and Dudley 2014, Saeed et al. 2015). White fringetrees may not act as a reservoir for EAB in part because these trees are generally small and cannot support many individual beetles. However, wild populations may support more EAB because these plants grow more abundantly in the wild compared those planted ornamental sites (Cipollini, personal observation), which could lead to increased use and higher beetle densities sustained by larger populations of white fringetrees. Emerald ash borer primarily using white fringetree during times of high beetle pressure (Peterson and Cipollini 2017) suggests that these trees will continue to get attacked at low rates throughout its native range and continue to experience low mortality. Some white fringetrees, however, can be killed or at least damaged severely enough by EAB to warrant removal in ornamental landscapes. This was the case for about 7% of the trees that we examined between 2015 and 2018. Because white fringetrees varied from being never attacked at all to severely enough to be killed, significant variation in the susceptibility to EAB attack or its response to it apparently exists in this species. Every location that we examined had at least some infested white fringetrees, thus attack was not geographically structured, and location cannot explain the variation in infestation statuses among trees. The phenotypic and genotypic basis of this variation needs further study. According to the USDA plant database, there used to be a coastal variety of white fringetree given subspecies status (Chionanthus virginicus L. var. maritimus Pursh) that has since been reclassified simply as white fringetree; therefore, other scientists have recognized morphological variation within this species (USDA NRCS 2019). A few varieties of white fringetree are currently offered in the horticultural trade (Niemiera 2010, Russ 2010), but they are uncommon and the majority of the trees are simply propagated from seed. If there is evidence of variation in plant resistance among white fringetrees, then those individuals or populations could be further investigated to understand potential mechanisms of resistance. We found EAB attacking fringetrees in several new locations and in several new states since our initial observations in 2015 (Peterson and Cipollini 2017). If more white fringetrees were systematically surveyed where EAB is already present, we would likely find additional trees under attack. As the invasion wave of EAB expands farther in North America, more white fringetrees will be exposed to attack. Importantly, the beetle is currently expanding in the native range of white fringetree where there are denser, wild populations and the tree is planted more commonly (USDA NRCS 2019). Emerald ash borer will have a continually increasing opportunity to infest white fringetrees in these locations. On the other hand, we have found that EAB is continuing to attack white fringetrees trees in areas, such as Ann Arbor, Michigan, close to where this beetle was first found in North America (USDA APHIS PPQ). White fringetrees in this area were likely among the first individuals of this species that these beetles ever encountered. White fringetrees in Ann Arbor had poor health ratings and high mortality rates from EAB attack because of their lengthy exposure to attack and some have been severely pruned or removed (Cipollini, personal observation). It is possible that EAB may be able to use white fringetree as a refuge to avoid biological control agents, such as Tetrastichus planipennisi, that has been released in the area to keep beetle densities below replacement (Duan et al. 2017). In controlled studies, however, T. planipennisi parasitized EAB larvae in white fringetree stems (Hoban et al. 2018, Olson and Rieske 2019). For white fringetree to serve as a reservoir for EAB, larvae will need to survive in densities high enough to overcome pressures from natural enemies such as T. planipennisi. One trait that may help white fringetrees tolerate EAB attack is due to having multiple stems; these trees are more resilient to attack than ash because they can lose several stems before being killed (Cipollini and Peterson, personal observations). Conversely, ash trees have one main stem, and once that stem is effectively girdled by EAB, they are more likely to succumb to attack. In conclusion, white fringetree is a novel host of EAB and will continue to get attacked throughout the United States, especially as the EAB invasion encroaches on the native range of this species in the southeastern United States. We anticipate finding variation in white fringetree resistance to EAB which could be from phenotypic or genotypic variation. Future research in plant resistance will help us better understand the variation in resilience that this tree has displayed over the course of 3 yr in our initial and follow-up studies. While this plant has suffered low mortality and appears to be a poor reservoir for EAB in low density plantings, white fringetree may serve as a better refuge in locations where it is more abundant, which could then support sufficient numbers of EAB to maintain a local population. This is most likely to occur in the heart of the native range of white fringetree in the southeastern United States where this tree grows most densely in the wild. Acknowledgments Thanks to Morton Arboretum, Chicago Botanic Garden, Purdue University, Indianapolis Museum of Art, Spring Grove Cemetery and Arboretum, Ferncliff Cemetery, Cox Arboretum, City of Dayton and Yellow Springs, Wright-Patterson Airforce Base, Wilmington College, and Hershey Gardens for access to white fringetrees. 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TI - The Fate of Ornamental White Fringetree Through the Invasion Wave of Emerald Ash Borer and Implications for Novel Host Use by This Beetle
JO - Environmental Entomology
DO - 10.1093/ee/nvaa018
DA - 2020-04-14
UR - https://www.deepdyve.com/lp/oxford-university-press/the-fate-of-ornamental-white-fringetree-through-the-invasion-wave-of-GaFfeiQgi0
SP - 489
VL - 49
IS - 2
DP - DeepDyve
ER -