Ailanthus
Suggested Citation: Huebner, Cynthia D.; Wickert, Kristen. 2025. Ailanthus (Ailanthus altissima). In: McNulty, Steven, project lead. The Silvics of North America. U.S. Department of Agriculture, Forest Service; Natural Resources Canada, Canadian Forest Service; National Forestry Commission of Mexico, collaborators. Washington, DC: U.S. Department of Agriculture, Forest Service. https://doi.org/10.2737/sna.ailanthus.altissima.
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- Steve McNulty, Project Lead (USDA Forest Service)
- Susan Iott, U.S. Project Coordinator (Unified Business Solutions, LLC)
- Sébastien Meunier, Canadian Project Coordinator (Canadian Forest Service)
- Cynthia F. Moser, Managing Editor (Unified Business Solutions, LLC)
Ailanthus altissima is identified as a widespread invasive nonnative species in North America. Management and utilization of this species should be carried out with caution in accordance with local, regional, and national regulations. Please note the following classifications present at the time of publication:
- Federal Government Documents
- United States of America: Listed as an introduced and widespread invasive (Category E) by the U.S. Geological Survey (USGS) “United States Register of Introduced and Invasive Species” (US-RIIS; Simpson et al., 2022)
- Canada: None available at this time
- Mexico: None available at this time
- National (nongovernmental) Assessments
- Reported to be invasive in natural areas (Swearingen and Bargeron, 2016)
Please see the following links to regional, State, and local listings highlighting invasive nonnative classifications:
- Invasive Plant Atlas of the United States: https://www.invasive.org/browse/subinfo.cfm?sub=3003
- USDA Natural Resources Conservation Service PLANTS Database: https://plants.usda.gov/plant-profile/AIAL/noxious-invasive
Pesticide Disclaimer
This publication/database reports research involving pesticides. It does not contain recommendations for their use, nor does it imply that the uses discussed here have been registered. All uses of pesticides must be registered by appropriate State and/or Federal agencies before they can be recommended.
CAUTION: Pesticides can be injurious to humans, domestic animals, desirable plants, and fish or other wildlife—if they are not handled or applied properly. Use all pesticides selectively and carefully. Follow recommended practices for the disposal of surplus pesticides and pesticide containers.
Ailanthus (Ailanthus altissima; figs. 1, 2) is a nonnative, invasive deciduous tree, originating from China, Taiwan, and northern Vietnam (Kowarik and Säumel, 2007). It is often associated with disturbed sites throughout much of its invasive range and is naturalized in Africa, Asia (beyond its native range), Europe, North America, and South America. Because it was cultivated as a horticultural, medicinal, or timber tree, its range may also have expanded beyond its native boundary within China (Hu, 1979). Its first record of introduction to North America was in 1784 in Philadelphia, Pennsylvania. This introduction came indirectly from trees produced from seed collected 30 years earlier in China by Pierre d’Incarville, who sent seed to colleagues in England in 1751 (Hu, 1979), misidentifying it as a varnish tree (Swingle, 1916). During the late 1800s, Chinese immigrants entering the western United States for railroad or mining work planted ailanthus (Feret, 1985). Ailanthus is notable for the stench that emanates from its crushed leaves; one of its modern common names in its native range is translated as “stinking tree.” An older common name in its native range is Shen shu or “God’s tree,” which may be the origin of “tree of heaven” (Swingle, 1916). Its species epithet, “altissima,” means “tallest” in Latin, referring to the tall heights this tree can reach (12 to 30 m [39 to 98 feet] at maturity; Dirr, 1998).
Figure 1—(A) Clonal growth of young ailanthus saplings and (B) enlarged view of the bark of a mature ailanthus tree. USDA Forest Service photos by Kristen Wickert.
Figure 2—(A) Male and female ailanthus small trees growing next to each other. USDA PLANTS Database photo by Douglas Goldman. (B) Enlarged view of leaf and leaflets of ailanthus; note glands at lower shallow leaf lobes. USDA Forest Service photo by Kristen Wickert.
In North America, ailanthus is found in a wide range of climatic conditions (Miller, 1990), elevations (Feret, 1985), and land types from open urban areas to closed-canopy forests (Hamerlynck, 2001; Hu, 1979; Knapp and Canham, 2000). This dioecious species grows and reproduces both vegetatively and sexually in rich and impoverished soils and under other stressful conditions (Marshall and Furnier, 1981; Pan and Bassuk, 1986), serving as a key metaphor in B. Smith’s 1943 novel, “A Tree Grows in Brooklyn.” Although relatively short-lived, ailanthus trees older than 100 years are not unusual in North America (Wickert et al., 2017) with some in Europe documented to be 130 years (Lauche, 1936). Clonal sprouts still existed as of 2001 from the first tree introduced in Philadelphia (Call, 2002), suggesting long-lived vegetative reproduction (over 200 years) for some ailanthus individuals.
Distribution
Ailanthus’ eastern range in North America extends from Vermont and the southernmost part of Ontario to as far south as Georgia; centrally it extends as far north as Wisconsin and as far south as southern Texas. Its distribution also extends as far west as southern British Columbia and southern California (fig. 3). This distribution as well as one using USDA Forest Service Forest Inventory and Analysis data located along the extent of the Appalachian Trail (Clark, 2013) highlights a lack of tolerance for extreme cold conditions. The Global Biodiversity Information Facility (GBIF), the Centre for Agriculture and Biosciences International (CABI), the Global Invasive Species Database (GISD), eFloraMEX, and the EPPO Global Database show ailanthus as naturalized in Mexico, but detailed location information is lacking (Walker et al., 2017).
Figure 3—Current distribution of ailanthus using point data from EDDMapS, the Global Biodiversity Information Facility (GBIF.org, 2024), and national forest inventory data from the United States (USDA Forest Service Forest Inventory and Analysis) and Canada (National Forest Inventory). USDA Forest Service cartography by Jacob Fraser.
Under intermediate (a representative concentration pathway, RCP, of 4.5) and extreme (RCP 8.5) climate scenarios of carbon dioxide concentrations, ailanthus is expected to gain a net 49 percent or 107 percent in area across North America, using Integrated Digitized Biocollections (IDigBio) and GBIF data, respectively, and WorldClim V1 variables Bio1 through Bio19 (Wang et al., 2022). Isler et al.’s (2023) management models for Switzerland predict that ailanthus will not displace most forest trees 200 or more years from now except potentially in drought-prone sites under extreme climate change (RCP 8.5). Models using distribution data from its native China underpredict its occurrence in the United States, suggesting this species has already expanded beyond what would be predicted from its native range (Albright et al., 2010).
Environmental Associations
Preferred climatic conditions of ailanthus range from temperate to subtropical and mesic to xeric (table 1). Ailanthus can tolerate xeric regions bordering the U.S. Great Plains with rainfall between 360 and 610 mm (14 and 24 inches) annually (Goor and Barney, 1968). Ailanthus can also tolerate the more mesic conditions in the southern Appalachian Mountains where rainfall may exceed 2290 mm (90 inches) annually (Frankson and Kunkel, 2022; Patterson, 1976). The annual minimum and maximum temperatures (calculated from monthly averages) for ailanthus are -9 and 36 °C (16 and 97 °F), respectively (Goor and Barney, 1968). Limitations of spread have been documented in the Mediterranean region at 11 °C (52 °F) and at elevations above 900 m (3,000 feet) above sea level. (Motti et al., 2021). Extreme cold and prolonged snow cover may restrict its elevational range to the lower slopes of the Rocky Mountains (Goor and Barney, 1968). Prolonged cold temperatures can kill seedlings and cause dieback in saplings and young trees, but resprouting may occur. Older trees may survive extreme cold weather (as low as -33 °C [-27 °F] over multiple days) but with subsequent reduced growth (Kowarik and Säumel, 2007).
| Summary statistics | Mean annual temperature in °C (°F) | Mean annual precipitation in millimeters (inches) | Elevation in meters (feet) |
| Minimum | 2.9 (32) | 126 (5) | Sea level |
| Lower 25 percent | 10.5 (37) | 983 (39) | 180 (580) |
| Median | 12.8 (39) | 1174 (46) | 272 (890) |
| Mean | 12.9 (39) | 1097 (53) | 414 (1,360) |
| Upper 25 percent | 15.1 (42) | 1304 (51) | 432 (1,420) |
| Maximum | 22.4 (50) | 2923 (115) | 3175 (10,410) |
Ailanthus is found in many substrates from barren, rocky sites to loams with clay or sand. It is also found in both calcareous and acidic soils that may be deep or shallow. It tolerates saline conditions (Dirr, 1976; Kowarik and Säumel, 2007; Singh et al., 1992). Ailanthus can tolerate pH as low as 4.1, phosphorus levels as low as 1.8 ppm, and saline conditions with salt levels as high as 0.25 mS cm-1 (electrical conductivity or millisiemens cm-1, which equals about 125 ppm) (Filippou et al., 2014; Plass, 1975). While ailanthus tolerates nutrient poor soils, it grows larger and faster in response to fertilizer with nitrogen, phosphorus, and potassium (Sojàk and Löffler, 1988). Soil compaction will reduce lateral and tap root growth but does not prevent ailanthus from becoming established (Pan and Bassuk, 1985).
The fossil record indicates that the genus Ailanthus was in western North America as far back as the early Eocene in the Cenozoic Era or about 55 million years ago. These records indicate that this genus, notably A. confucii, which most resembles A. altissima, was circumboreal. The Ailanthus genus went extinct in North America during the cooling of the Pleistocene (Corbett and Manchester, 2004).
Sexual Reproduction
Ailanthus is predominantly dioecious. Hermaphroditic and monoecious individuals have been documented, but they are few in number (Dirr, 1998). Inflorescences are shaped like a thyrse (a raceme with cymose lateral branches). Male inflorescences are larger in size with more flowers than female inflorescences; flowers of both sexes are small and yellowish green in color (figs. 4A, 5). The stamens in the female flowers are vestigial (i.e., they produce no pollen) and the male flowers have 10 fully functional stamens with no pistils (Hu, 1979). Flowering usually starts when the plant is between 3 and 5 years of age, though plants have been known to flower as early as 3 to 6 weeks after germination, especially if exposed to high light and extended photoperiods (Bory and Clair-Maczulajtys, 1977; Feret, 1973), but the seeds are not likely to be viable (Bory and Clair-Maczulajitys, 1977). Flowering starts from mid-May or as late as early July (Castro-Diéz et al., 2014). Male trees often start flowering a few days earlier than female trees within a local area and emit an unpleasant odor (Bory and Clair-Maczulajtys, 1982).
Figure 4—(A) Male inflorescences and (B) immature samaras of ailanthus. USDA PLANTS Database photo by Douglas Goldman.
Ailanthus is pollinated by insects (including ants, bees, beetles, and flies; Aldrich et al., 2008) and wind (Maxia and Maxia, 2003). Insects are attracted to either the floral or extrafloral nectaries or the pollen (or combination thereof). Floral nectaries are found on the glandular discs of both the male and female flowers (Bory and Clair-Maczulajtys, 1982). Extrafloral nectaries are found at the apex of shallow leaf lobes, pseudostipules (modified leaflets at the base of each leaf petiole), and cataphylls (e.g., bud scales) (Davies, 1946).
Fruit set begins in early June and lasts until late October (Castro-Diéz et al., 2014). Fruit are spirally twisted samaras with a centrally located seed; one to five seeds can develop from the five carpels (fig. 4B). Seeds do not contain endosperm, but cotyledons have stored oils that may allow for some seed longevity (Little, 1974). Samara size, color, weight, and thickness vary considerably in China (Xiong et al., 1983) and Europe (Bory and Clair-Maczulajtys, 1980) but have less variation in North America (Feret and Bryant,1974; Feret et al., 1974). Seed size and mass in North American populations increase with latitude (Feret et al., 1974).
Figure 5—(A) Male and (B) female flowers of ailanthus. USDA PLANTS Database photo by Douglas Goldman.
Ailanthus’ phenology is characterized by late leaf flush and flowering in spring (Castro-Diez et al.,2014) and an early leaf fall in autumn compared to associated deciduous native species (e.g., ash [Fraxinus]). Ailanthus buds fail in response to late frosts (Davies and Theiss, 1937; Theiss, 1933). After leaf fall in early November in one study in Indiana, buds required a minimum of 41 days at temperatures around 25 °C (77 °F) and adequate moisture to force bud-break on cut stems, with some buds never opening. This may indicate a high forcing need at low chilling temperatures experienced in November. The duration of the forcing period decreased with increasing time since leaf fall, reaching zero days by early March (Davies and Theiss, 1937). Foliar application of abscisic acid (ABA) triggers growth cessation in ailanthus (Lavender and Silim, 1987), indicating that day length and decreasing temperatures may also play a role. Ailanthus’ short growing season could be adaptive at low elevations to reduce the risk of frost damage in the spring but could also limit its ability to colonize cold climates because a shorter growing season in such climates would result in a negative carbon budget (Motti et al., 2021).
There may be as many as 500 seeds per inflorescence and one 3-m-tall tree may produce as many as 650 clusters, for a total of 325,000 seeds per tree each growing season (Bory and Clair-Maczulajtys, 1980) or an estimated 10 to 50 million seeds over a tree’s lifetime, depending on tree age and site conditions (Wickert et al., 2017). Samaras travel as far as 100 m (330 feet), more rarely up to 200 m (660 feet) (Landenberger et al., 2007) by wind, and possibly longer distances by water (Kaproth and McGraw, 2008; Kowarik and Säumel, 2007), although they often stay on the trees through winter and into the next spring. The degree of spiral-twisting determines the samara’s ability to autorotate, increasing the dispersal distance (Planchuelo et al., 2017). Rodents utilize the samaras to pad nests, theoretically dispersing and “planting” ailanthus (Bory and Clair-Maczulajtys, 1980). Vehicles or machinery may also disperse seeds. Seedbanks are less likely to form under suitable germination conditions due to a high germination rate that may reach 98 percent. However, dormancy has been documented (Redwood et al., 2019) and seedbanks have been noted in a New York urban forest (Kostel-Hughes et al., 1998). Seed longevity has also been found to be as long as 5 years in the laboratory or when seeds are buried 10 cm (4 inches) in field soil. Nonetheless, 60 to 70 percent of the germination occurs within the first year for seeds just under the litter layer (Rebbeck and Jolliff, 2018). Seed predation also occurs in the nonnative range; some bird species (Vines, 1977), deer, mice, and voles are known to feed on the seeds (Cadenasso and Pickett, 2000; Facelli, 1994; Ostfeld et al., 1997).
Ailanthus does not require seed pretreatment for germination to occur, but cold stratification for 30 to 60 days improves the germination rate (Zasada and Little, 2008). Germination rates are greater in sites with higher light levels (Huebner et al., 2018), but this is likely due to the higher soil temperatures associated with those sites. Optimal germination rates occur around 30 °C (86 °F) with lower rates documented at 25 and 40 °C (77 and 104 °F) (Kheloufi et al., 2020); light is not a requirement for germination. Lower temperatures (6 to 15 °C [43 to 59 °F]) appear to promote dormancy in ailanthus (Pepe et al., 2020). Heat ranges typically associated with prescribed burning (80 to 110 °C [176 to 230 °F] for 5 to 10 minutes) do not stimulate germination in ailanthus and temperatures above 110 °C kill the seeds (Cruz et al., 2021). Seeds and seedlings do not exhibit salt tolerance and are sensitive to salt stress (Kheloufi et al., 2020). However, Bicknell and Smith (1975) found only a minimal reduction in ailanthus seed germination at salt concentrations typical of roadsides with drainage from salted roadways.
Asexual Reproduction
Ailanthus can reproduce asexually from lateral roots as suckers, stumps (originating from stem or root tissue), and root fragments, and from stem fragments via adventitious roots (Kowarik and Säumel, 2007). Root segments as small as 1 cm (0.4 inch) may produce shoots due to the many suppressed buds or primordia found on the lateral roots (Inverso and Bellani, 1991). Seedlings cut when less than 1 year old may sprout and produce root suckers (Bory et al., 1991). Two-year-old ailanthus seedlings may produce lateral roots as long as 2 m (7 feet) (Pan and Bassuk, 1986) and adult trees have documented lateral roots as long as 27 m (89 feet) (Kiermeier, 1987 as interpreted by Sladonja et al., 2015) and tap roots as deep as 2 m (Singh et al., 1992). Clonal expansion of individual genotypes appears to be limited to about 45 m (150 feet) (Chuman et al., 2015). Clonal growth results in a patchy spatial distribution of male and female clones either along a disturbed corridor or within a broader landscape. Larger patches up to 120 m (394 feet) long have been observed (Kowarik, 1983 as interpreted by Sladonja et al., 2015), but these are likely to be composed of more than one genotype (Chuman et al., 2015). In areas of dense ailanthus stumps, asexual reproduction dominates over sexual reproduction, with more sprouts and suckers than seedlings. Sprouts and suckers generally grow faster and are larger than seedlings; sprouts and suckers could be as much as two or three times the height, respectively, as seedlings in the same time period (Yonebayashi et al., 2016). Sprouts and suckers may also start reproducing earlier and grow larger over the same timespan and, thus, produce more seed than trees originating from seed, but studies have not been conducted yet to confirm this hypothesis.
Though not a requirement for vegetative growth (Kowarik, 1995), stem damage from cutting, wind, frost, herbivory, and disease as well as top-kill due to fire or inadequate herbicide application will stimulate sprouting and root suckering (Bory et al., 1991; Yonebayashi et al., 2016). Herbicide efficacy is greater in the fall when photosynthates are being transported to the root system (European and Mediterranean Plant Protection Organization, 2020). Consequently, sprouting due to any form of damage may be more severe in late spring and early summer when transport of photosynthates is primarily toward the shoots.
Ailanthus is associated with temperate, broadleaf, deciduous forests as a pioneer species in its native range. In its invasive range, it is most closely associated with urban and disturbed natural areas but may also be found in closed canopy forests in temperate, Mediterranean, and subtropical plant associations. Though characterized as a shade-intolerant tree, sprouts exhibit greater shade-tolerance than seedlings (Knapp and Canham, 2000; Kowarik, 1995) and ailanthus has exhibited a degree of shade tolerance as a sapling that is not shown for other shade-intolerant species such as black locust (Robinia pseudoacacia) (Vítková et al., 2017), staghorn sumac (Rhus typhina) (Tan et al., 2018; Zhang et al., 2009), Scots pine (Pinus sylvestris), and European larch (Larix decidua) (Knüsel et al., 2017) but is similar to gap-obligate deciduous species such as tuliptree (Liriodendron tulipifera) (Knapp and Canham, 2000). Ailanthus is documented as showing a shared patchy distribution pattern (i.e., found in close proximity to each other relative to other species) with species that have an affinity for similar site conditions, most notably black locust (Call and Nilsen, 2003; Nilsen and Huebner, 2023). It is more likely to prosper in sites with high soil fertility, reliable precipitation, and good drainage, but it is tolerant of more xeric conditions as well as low soil fertility. It may be among the few plant species that can become established, grow, and reproduce in some of the most impoverished sites (Marshall and Furnier, 1981; Pan and Bassuk, 1986). Both the native and nonnative ranges of ailanthus are characterized by long and warm growing seasons, non-extreme winter frosts and annual precipitation of more than 500 mm (20 inches) (Kowarik and Säumel, 2007) associated with temperate climates.
Ailanthus is found in all successional stages (e.g., old fields to mature forests) and may function as a subcanopy tree or a dominant canopy tree depending on the site history and species composition in its nonnative range (Nilsen and Huebner, 2023). Ailanthus tends to have the largest percentage of stems (about 72 percent) in the smallest size class, suggesting a high mortality rate among young saplings. Nonetheless, transition probabilities from old field to a system dominated by shrubs and then trees in its nonnative range are relatively high, estimated at 20 percent (shrubs) and 25 percent (trees), compared to 4 percent (shrub) and 20 percent (trees) of other species (Mücke and Kliese, 1991 as interpreted by Kowarik and Säumel, 2007). Its positive association with black locust is maintained through successional stages, but ailanthus may have an inhibitory effect not shared by black locust on the regeneration of other associated native species, possibly due in part to its allelopathic properties (Nilsen and Huebner, 2023).
Seedlings of ailanthus can grow 1 to 2 m (3 to 7 feet) in height in a single year (Hu, 1979) and it may be the fastest-growing tree species in North America (Howard, 2004; Knapp and Canham, 2000) and Britain (Mabberley, 1997). Incremental height and diameter growth are largest in trees that are 5 to 10 years of age. If site conditions remain favorable, these rapid growth rates may continue to an age of 10 to 20 years. Growth then tends to decrease after 20 years of age. For example, in Italy trees are documented as growing 5 to 10 mm year-1 (0.20 to 0.39 inch year-1) in diameter in trees 5 to 10 years of age and 3 to 4 mm year-1 (0.12 to 0.16 inch year-1) for 50-year-old trees (Speranzini, 1937). These growth rates are significantly higher than associated native trees regenerating in forest gaps in the northeastern United States (Knapp and Canham, 2000). Growth rates are more rapid in more open, urban areas (Pan and Bassuk, 1986) than in forests (Kota et al., 2007). Ailanthus exhibits higher growth rates under warmer and drier conditions (Knüsel et al., 2019b). Though ailanthus exhibits a degree of shade tolerance allowing it to survive and grow under closed canopy conditions, it manifests more rapid growth under higher light (Knüsel et al., 2017). The mechanism behind its survival and growth under a canopy may be ailanthus’ tendency to increase its specific leaf area in response to shade (Petruzzellis et al., 2019). Ailanthus has been labeled a low-quality timber species, and larger ailanthus trees (>75 cm [30 inches] in diameter) may be prone to heart rot (Kasson et al., 2013a). Timber yield estimates are lacking (Miller, 1990), though the potential for high yields has been described (Brandner and Schikhofer, 2010). Pulp yield for papermaking has been estimated at 49 percent (weight in weight, or w/w; in 100 g of total processed wood material, 49 g is pulp) from 2-year-old trees; older trees have a lower pulp yield (Baptisa et al., 2014).
Management with Natural Regeneration
Ailanthus responds positively with increased growth in most silvicultural systems that include opening the canopy or disturbing the soil (or both). If germinating from seed, greater canopy opening will help ensure warmer soil and greater germination rates. Light levels of 15 percent photosynthetically active radiation (PAR) or above will help ensure survival and more rapid growth of seedlings (Huebner et al., 2018). If the trees being harvested are ailanthus and there is no herbicide application to the stumps, the site is likely to be dominated by a 1.6- and 26-fold increase in sprouts and suckers, respectively, per cut stem (Burch and Zedaker, 2003; Kowarik and Säumel, 2007). Silvicultural practices may be one of the main drivers of spread of ailanthus (Erdélyi et al., 2021).
Scenario 1: Ailanthus not present in the stand. Pre-cut disturbances to prepare a site may promote pre-cut invasions. Removal of aggressive nonnative or native vines and trees may help ensure an extended period of canopy openness, which may promote germination of any ailanthus seed dispersed from nearby sites. Removal of dominant shrubs and forbs in the understory with herbicide or fire may help ensure germination of any ailanthus seed that is dispersed from adjacent sites (Huebner et al., 2018; Radtke et al., 2013).
Scenario 2: Ailanthus present as seedlings, saplings, or reproducing trees (or combination thereof) within the stand with no preharvest removal from the stand. Removal of other aggressive nonnative or native vines and trees may ensure increased germination and increased growth rates of the seedlings, saplings, and young ailanthus trees in response to increasing soil temperature and light. Removal of dominant understory shrubs and forbs using fire or herbicide (or both) may also promote establishment of new ailanthus seedlings, sprouts, and suckers (Huebner et al., 2018; Radtke et al., 2013).
Scenario 3: Ailanthus and other invasive nonnatives as well as interfering native species removed and treated prior to stand harvest. Removing all ailanthus seedlings, saplings and trees with herbicide to ensure a complete belowground kill will reduce any negative impacts it may have on the desired native tree species. Removal of dominant understory species will also reduce competition with the desired native tree species. See the section Silvicultural Options and Consideration – Scenario 3 for details on use of herbicide on ailanthus.
Management with Planted Regeneration
These strategies are relevant only if this invasive species is grown in a plantation. If maintaining genetic diversity or specific genotypes of ailanthus is a goal, the planting of seedlings should be done with an approximate free radius of 45 m for each seedling or sapling. Each planted seedling or sapling, especially if any damage occurs to the stem, is likely to sprout and sucker up to this distance (possibly further) from the seedling or sapling (Chuman et al., 2015). Planting may allow a land manager to ensure that particular genotypes (possibly for those most suitable for drier conditions or for biofuels if such genotypes or populations are ever determined to exist) are grown.
The same site preparation described in the section Natural Regeneration would apply here.
Tending or Intermediate Management
With lower soil temperatures, germination may also decrease (Huebner et al., 2018; Kheloufi et al., 2020). Periodic opening of the canopy, subcanopy, and understory with thinning or removal of competing species will ensure more rapid growth of ailanthus as seedlings, saplings, sprouts, and suckers (Radtke et al., 2013).
Thinning of ailanthus stands without killing the stumps of the removed ailanthus trees with herbicide will result in sprouting and suckering. If thinned trees are killed, remaining ailanthus stems will be released and increase in size. Depending on the amount of canopy opening, the number of ailanthus seedlings may increase due to outside seed sources or seed production from the remaining, larger trees (see preceding sections). However, the larger ailanthus trees may also reduce the amount of canopy opening, resulting in a reduction in ailanthus seed germination and sprouting and suckering from poles and trees. Poles and trees are likely to slow their growth as the stand ages (Speranzini, 1937).
The sympodial modular growth of ailanthus, which gives rise to a spirally radiating branch pattern and young trees that are nearly as wide as they are tall (Kowarik and Säumel, 2007), may make pruning desirable. However, doing so is likely to stimulate sprouting or root suckering and is not advised, unless this is a desirable outcome.
Unless sprouting from stem segments is desirable, all slash should be removed from any site with ailanthus removal treatments. Allelopathic properties of leaves and wood may result in legacy effects on soil with decomposing slash. This may vary with soil pH. The allelopathic compounds, ailanthone and quassin, break down in alkaline soils (pH = 10), whereas they do not in acidic or sterile soils, suggesting a microbial mechanism behind the loss of allelopathic compounds (Sasnow, 2012). However, leaving slash behind may decrease light and soil temperature, possibly inhibiting some establishment of ailanthus seedlings, sprouts and suckers. Leaving slash of other species also serves as a deer deterrent (Kota and Bartos, 2010; Puettmann et al., 2023) that may foster growth of native trees more prone to deer herbivory.
Ailanthus poles and trees that are removed from a site can be used for fuel, paper, furniture, biofuel, and charcoal (Baptista et al., 2014; Fry, 2010; Panayotov et al., 2011; Shah et al., 2014; Terzopoulou et al., 2023). Myocarditis (minor to death) from exposure to ailanthus’ sap is documented, as is contact dermatitis (Bisognano et al., 2005; Braunstein and Polis, 2005; Whiticar and Harvey, 2009). Consequently, care should be taken when removing this tree to limit exposure to the sap.
The genus Ailanthus globally has six accepted species, all with similar morphology with no documented hybridization. There are no currently recognized subspecies of Ailanthus altissima (Mill.) Swingle, but there are currently three accepted intraspecific names: Ailanthus altissima var. altissima (Mill.) Swingle, A. altissima var. sutchuenensis (Dode) Rehd. & Wilson, and A. altissima var. tanakae (Hayata) Kanehira et Sasaki. These varieties differ in samara size, with the sutchuenensis variety having the largest samaras and the tanakae variety having the smallest (Hua and Thomas, 1997). Ailanthus altissima var altissima is the variety that has invaded North America and other continents. There are 17 known synonyms with the earliest being Rhus sinense Ellis (1757) and the latest Ailanthus peregrina (Buc’oz) F.A. Barkley (1937) (Kowarick and Säumel, 2007). There are at least seven ailanthus cultivars (Kowarick and Säumel, 2007). Plants with reddish/pink fruits are combined as A. altissima f. rubra (Dippel) Geerinck comb. nov. (Geerinck, 2002). Phylogenetic analyses using chloroplast genome sequencing place A. altissima and its family Simaroubaceae most closely related to the Rutaceae family (Saina et al., 2018). The first seed brought to North America is most likely haplotype 7 (Neophytou et al., 2020) from the central-eastern provinces of Anhui and Jiangsu (northwest of Shanghai, China) and areas north of Beijing, China (Liao et al., 2014).
Genetic Variation
The within-population genetic diversity (He) of ailanthus in its native range is 0.547 to 0.772, with 1 being the highest diversity (Saina et al., 2023; Zhang et al., 2023). Studies in ailanthus’ invaded range in Europe (He = 0.51 to 0.60; Neophytou et al., 2020) and North America (He = 0.46 to 0.69; Aldrich et al., 2010) show moderate genetic diversity. There are significant genetic differences among populations that support multiple introductions and subsequent gene exchange among populations as well as bottlenecks due to founder effects from small starting populations (Aldrich et al., 2010; Neophytou et al., 2020). Higher genetic diversity of ailanthus in its nonnative range was found in the midwestern United States due to intermixing of haplotypes from the west and east coasts. Higher genetic diversity for ailanthus in its nonnative range is also likely to be found in or near larger cities, a result of long-distance dispersal along major roadways (Aldrich et al., 2010). Brusa and Holzapfel (2018) found that genetic diversity increased with degree of urbanization, with gene flow being related to the level of human traffic to a site. Managed forests in North America, however, reduced admixture, suggesting that managing forests for ailanthus removal can reduce gene flow (Brusa and Holzapfel, 2018).
Most ailanthus individuals are diploid with 64 chromosomes (Ivanova et al., 2006) with some reports of individuals having 80 chromosomes (Desai,1960). A triploid individual, documented in Japan, had abnormally large pollen that may reduce fertility. If a reduction in fertility can be confirmed, triploid ailanthus could be cloned as a cultivar with less invasive tendencies (Kurokochi et al., 2014).
Ailanthus has been successfully transferred unintentionally in its nonnative range along major roadways between urban areas (Aldrich et al., 2010; Brusa and Holzapfel, 2018). It was intentionally introduced multiple times directly from its native range and indirectly from the United Kingdom as a horticultural tree (Hu, 1979). Intentional introductions may still occur in North America but are currently discouraged or regulated against. However, one can still purchase (as of 2023) ailanthus bare-root seedlings and seed online.
There has been breeding for the red fruit trait in China (Ma et al., 2023), wood for better biofuel in its native and nonnative range (Kashe et al., 2021; Terzopoulou and Kamperidou, 2022), and silkworm in its native and nonnative range (Eshankulov and Xolova, 2020).
Genomic Resources for Forest Tree Species
Among the first genetic studies on ailanthus, Dallas et al. (2005) developed nine nuclear microsatellites for ailanthus, which can be found in GenBank by searching each accession number (https://www.ncbi.nlm.nih.gov/nuccore/). Neophytou et al. (2018) developed 19 additional nuclear microsatellites from which next-generation sequencing allowed the discovery of single nucleotide polymorphism (SNP) information although no SNP arrays are currently listed for ailanthus (GenBank; https://www.ncbi.nlm.nih.gov/nuccore/).
The genome sequence is 939 megabases in span and is stored in GenBank under the Taxonomy ID: 2768810. Most of the assembly is scaffolded into 31 chromosomal pseudomolecules. The mitochondrial and plastid genome assemblies are 661.1 kilobases and 161.1 kilobases long, respectively (Schley et al., 2023). Ailanthus’ chloroplast complete genome and several renditions of partial sequences have been sequenced and are available in GenBank (Saina et al., 2018).
There are no insects native to North America that rely solely on ailanthus for their food or habitat. Many insects and pathogens that interact with ailanthus also originate from East Asia, some of which are detrimental to native and agricultural plants (Wakie et al., 2020). For example, spotted lanternfly (Lycorma delicatula) (table 2; fig. 6) shows a preference for feeding and reproducing where ailanthus is abundant (Laveage, 2023; Murman et al., 2020; Uyi et al., 2021). The brown marmorated stink bug (Halyomorpha halys) and ambrosia beetle (Euwallacea validus; sometimes referred to as a shot-hole borer beetle; fig. 6B, E) are nonnative insects that feed on ailanthus and may also harm native forest and agricultural species (Aigner et al., 2017; Baker, 1972) (table 2).
| Degree of impact | Roots | Bole (bark, phloem, and xylem) | Foliage, shoot, and twig | Flowers, fruit, and seeds |
| Greatest impact | N/A | Wilt diseases: Verticillium nonalfalfaea—native pathogen causes mortality in as little as 3 months and Verticillium dahliaeb—native pathogen causes mortality in months to years. | Spotted lanternfly (Lycorma delicatula)c—nonnative hemipteran insect that feeds on ailanthus in all life stages. | N/A |
| Moderate impact | Honey mushrooms (Armillaria spp.)d— multiple species across the region acting as a secondary pathogen. | Snout weevil (Eucryptorrhynchus brandti)e—nonnative species being considered for biological control; not released as of 2023. Ambrosia beetle (Euwallacea validus)f—nonnative ambrosia beetle that infests stressed and dying ailanthus. | Ailanthus webworm (Atteva aurea)g—expanded from native range (Florida and south to Costa Rica) to across the United States. Larvae eat foliage and cover new growth in webbing. | N/A |
| Low impact | N/A | Agrilus smaragdifronsh—newly introduced buprestid beetle; likelihood of potential damage currently unknown, but beetle is spreading across the Mid-Atlantic region. There are several basidiomycotan fungi that utilize ailanthus as a food source and contribute to the decay of the dead tree, such as Schizophyllum commune, Ganoderma applanatum, Daedalea spp., Trametes versicolor, Exidia recisa, Auricularia spp., Pleurotus ostreatus, and Flavodon ambrosius.i | Asiatic garden beetle (Maladera formosae)j—adults feed on leaves. Ailanthus silkmoth (Samia cynthia)k—caterpillars feed on leaves; this insect was intentionally brought to the United States for sericulture but failed as a commercial source of silk and populations do not survive long in the wild.l Fall webworm (Hyphantria cunea), banded tussock moth (Halisidota tessellaris), and white-marked tussock moth (Orgyia leucostigma) are examples of native insects that utilize ailanthus but do minimal damage.g Several well-studied pathogenic genera of ascomycotan fungi attack ailanthus, causing foliar disease but rarely resulting in tree death. Genera include the following: Diplodia, Armillaria, Phymatotrichopsis, Diaporthe, Eutypella, and Collectotrichumg phytoplasms—reported in Europe as causing disease in ailanthus foliage, but symptoms have not been seen in North America.m Eriophyid mite (Aculus mosoniensis)—being studied in Europe as a possible biological control.n | Brown marmorated stink bug (Halyomorpha halys)o—all life stages of this insect feed on foliage and seeds and can introduce pathogens and necrosis spots. |
N/A: not applicable.
aFormerly called Verticillium albo-atrum. Currently being researched as a potential biological control for ailanthus. Sources: Inderbitzin et al. (2011), Schall et al. (2009).
bCauses mortality but after several years of moderate infection. Source: Pisuttu et al. (2023).
cSource: Hoover et al. (2023).
dSeveral species of Armillaria exist and are being further described in North America. It is not fully understood which of these regional species do or do not infect ailanthus. There are historical records of infection in the Mid-Atlantic region by Armillaria mellea. Source: Ding et al. (2006).
eThough not yet released in North America, this nonnative species has potential to utilize and damage ailanthus if introduced in North America and to be potentially used as a biocontrol. Sources: Herrik et al. (2012), Kok et al. (2008), McAvoy et al. (2014).
fSource: Cognato et al. (2015).
gSource : Ding et al. (2006).
hSource: Hoebeke et al. (2017).
iSources: Buzina et al. (2001), Charpentier-Alfaro et al. (2023), Kasson et al. (2016).
jSource: Baker (1972).
kSources: Baker (1972), Peigler and Naumann (2003).
lSource: Peigler and Naumann (2003).
mSource: Krstić et al. (2022).
nSource: Marini et al. (2021).
oSource: Acebes-Doria et al. (2016).
Ailanthus webworm (Atteva aurea) is a native insect expanding its range from southern Florida to most of ailanthus’ invasive range in North America. The larval stage of this insect feeds on the foliage but does not cause death in saplings and adult trees, though some seedlings may die from infestation (Ding et al., 2006).
Two native Verticillium species cause severe wilt disease of ailanthus (fig. 6C, D). Research into the question of utilizing Verticillium nonalfalfae as a biocontrol shows that this fungus can kill young trees in 3 months and larger trees in 1 year (Schall and Davis, 2009). Host range studies confirm that the fungus does not cause significant mortality to native plants in North America, though death and natural spread to some native plants is documented (Kasson et al., 2015; Pile Knapp et al., 2022; Schall and Davis, 2009). Naturally occurring ailanthus infections of V. nonalfalfae have been confirmed in Pennsylvania (Schall and Davis, 2009), Ohio (Rebbeck et al., 2013), and Virginia (Snyder et al., 2013).
Figure 6—(A) Spotted lanternfly (Lycorma delicatula); (B) Euwallacea validus, a nonnative ambrosia beetle; (C) varying stages of foliage wilt on ailanthus after Verticillium infection; (D) vascular discoloration of the xylem tissues of ailanthus after Verticillium infection; (E) the internal galleries of Euwallacea validus which harbor mycelial growth of a fungal symbiont, Fusarium. USDA Forest Service photos A, B, C, E by Kristen Wickert and photo B by Molly Sherlock.
Although ailanthus survives moderate insect infestations, spotted lanternfly can cause growth declines in ailanthus after 3 years of feeding (Hoover et al., 2023).
Verticillium wilt increases tyloses in ailanthus’ vascular tissue; the tyloses compartmentalize the disease but block water flow in the tree’s vascular system (Fradin and Thomma, 2006). This compartmentalization results in yellow streaking under the bark (Kasson et al., 2014). Verticillium wilt disease may also clog the vessels with fungal spores (Inderbitzin et al., 2011; Pisuttu et al., 2023; Schall and Davis, 2009).
Stressed ailanthus attracts the nonnative ambrosia beetle, which may transmit the Verticillium fungi from infected to healthy ailanthus trees (Aoki et al., 2018; Cognato et al., 2015; Kasson et al., 2013b; O’Donnell et al., 2015). The snout weevil (Eucryptorrhynchus brandti) can cause death of ailanthus in the tree’s native range, but primarily acts as a secondary interaction after the tree has been weakened by pollution, other insects and pathogens, or a lack of resource (McAvoy et al., 2014). Spotted lanternfly and other stressors, such as drought, may serve a similar function to ensure high mortality of ailanthus when V. nonalfalfae is present.
Use of Verticillium wilt as a biocontrol of ailanthus combined with other stressors and potential vectors of spread may help ensure greater ailanthus mortality. Some native trees often found in disturbed areas near ailanthus may manifest low mortality rates in response to the fungus, including black locust and staghorn sumac (Kasson et al., 2015).
Ailanthus regeneration is not associated with a known fire regime in its native range or North America. It is a thin-barked species with no evident fire adaptations except the ability to sprout and root sucker after a moderate or light burn (Maringer et al., 2012; Rebbeck et al., 2019). Ailanthus’ presence is not yet documented to increase the likelihood of fire although at dry sites, the large addition of leaf litter associated with ailanthus (Medina-Vellar et al., 2015) may add fuel and make a dry site more prone to fire. However, a thick, wet litter layer in more mesic sites may conversely reduce the likelihood of fire. One model developed for xeric pine-oak forests using LANDIS, a spatially explicit successional model, predicted that intermediate levels of fire frequency promoted the spread of ailanthus, especially in fragmented landscapes, while low and high fire frequencies reduced its spread while increasing the growth of associated native pine species (Xi et al., 2010).
Although there may be a reduction in stems from burning due to top-kill and some mortality after a burn, especially if severe, ailanthus tends to resprout and sucker after a burn, with sprouting and suckering generally decreasing with time since the burn (Maringer et al., 2012; Pomp, 2008; Rebbeck et al., 2019). Although germination is not stimulated directly by fire, conditions after a burn (increased light, decreased litter, decreased competition, increases in soil nutrients) are associated with an increase in ailanthus’ germination rate (Guthrie et al., 2016; Knüsel et al., 2019a; Yonebayashi et al., 2016).
Mortality or weakening due to pathogens, insect infestations or drought may make ailanthus trees more prone to fire by adding fuel to a site. Fire does not appear to impact ailanthus webworm either positively or negatively, unlike other moth species in which moth abundance increases in burned sites as basal area increases (Guerra et al., 2019). Fire after a drought or insect/pathogen outbreak may result in greater mortality of ailanthus (Fernandez-Manjarrés et al., 2018). Because fire tends to reduce the abundance of competing species, at least initially, this also tends to favor ailanthus abundance such that any biotic resistance that may be helping to suppress ailanthus growth and reproduction may be removed after a burn (Crandall and Knight, 2018).
Use of fire alone as a management tool may pose too high a risk in sites that are invaded by ailanthus or that are adjacent to significant ailanthus seed sources, unless fire intensity can be controlled (Maringer et al., 2012; Rebbeck et al., 2019; Xi et al., 2010). However, combining fire with a drought or pathogen outbreak may result in increased mortality of native species, resulting in increased invasion (Fernandez-Manjarrés et al., 2018) unless combined disturbance levels are intense or frequent enough to increase ailanthus mortality as well; such thresholds of intensity and frequency are unknown for ailanthus.
Ailanthus can be found in habitats limited by water, usually in conjunction with higher temperatures; it is not typically found in xeric, cold climates (Kowarik and Säumel, 2007). It has been used in its native range in China to restore landscapes impacted by severe droughts (Zhang et al., 2014). Ailanthus’ invasive range may be broader than what would be predicted by variables associated with its native distribution, including possible expansion into drier environments (Albright et al., 2010).
Ailanthus exhibits high hydraulic efficiency with wide xylem conduits, permeable interconduit pit membranes, and low wood density or construction costs. This hydraulic efficiency comes at the expense of hydraulic safety; ailanthus is vulnerable to xylem embolism (Petruzellis et al., 2019). Ailanthus is, however, able to close its stomata rapidly in response to severe drought as well as rapidly recover or reverse any embolism. Shrunken cortical cells and a multilayer endodermal-like tissue with suberized cells in the roots may in turn prevent water loss from the roots to the dry soil (Trifilò et al., 2004). Though there is evidence of drought tolerance, ailanthus may also exhibit a reduction in growth as a seedling in response to a lack of water; younger seedlings (e.g., less than 2 years) may be less drought tolerant than older seedlings (Stevens et al., 2018), or seedlings at higher elevations or colder climates may be less apt to close their stomata or shrink their cortical root cells in response to a lack of water. Older trees (43 to 62 years) in Switzerland show a lower reduction in growth rate during a severe drought compared to an associated native tree, suggesting ailanthus maintains a degree of drought tolerance into maturity (Knüsel et al., 2015). Plasticity of leaf mass per area (dry leaf mass divided by fresh leaf area) of ailanthus may enable this species to withstand relatively short periods of drought while maintaining water content and gas exchange during photosynthesis (Pepe et al., 2022). This plasticity and ability to survive relatively short drought periods also enable ailanthus to take advantage of periodic water availability, such as that found in irrigated xeric climates. However, although ailanthus can maintain high growth rates compared to other species at high temperatures that would also be found in these xeric environments, its growth slows in response to temperatures above 21 °C [70 °C] (Knüsel et al., 2019), possibly resulting in smaller xylem cells and reducing its capacity to recover from embolism events.
Drought conditions may make ailanthus more susceptible to pathogens, herbivory, and fire. Indeed, the wilt diseases caused by Verticillium species typically less deadly than V. nonalfalfae are likely exacerbated by drought and vice versa (Pisuttu et al., 2023), possibly increasing mortality rates associated with pathogens that without a drought would just be a nuisance.
Though ailanthus exhibits some drought tolerance, interactions with other disturbances such as fire, insects, and diseases may be additive or synergistic, resulting in increased mortality rates of ailanthus under these combined conditions for both mesic and xeric sites. If the sites invaded are in xeric sites where drought-tolerant and fire-adapted native species are present, ailanthus could decrease in abundance in response to extended dry conditions. This reduction in ailanthus may provide native species an opportunity to dominate once again. Unfortunately, in more mesic sites, increased drought conditions and ailanthus’ ability to recover from embolism may make those sites more susceptible to invasion and spread of ailanthus in response to a drought (Petruzellis et al., 2019).
Hurricanes, tornadoes, and windstorms have been documented to result in increases in several invasive plants, including ailanthus, in response to uprooting existing native trees (Culley et al., 2023; Daniels and Larson; 2019; Xi and Peet, 2008). Though described as not tolerant of flooding (Miller, 1990), ailanthus is found to spread in some floodplains that may or may not be seasonally dry (Castro-Díez et al., 2014; Liess and Drescher, 2008) and a freshwater tidal estuary in Maryland (Kiviat, 2004).
Ailanthus’ typical response to hurricanes, tornadoes, and windstorms has been prolific sprouting and suckering due to damaged stems, and prolific seed germination within the openings caused by these disturbances. Rapidly moving water in a flood may also result in damaged stems that lead to sprouting and suckering. If water recedes, both ailanthus sprouts and seedlings may take advantage of the open area. Flooding also serves a good vector, spreading ailanthus seed (Kaproth and McGraw, 2008) and possibly root and stem segments that may sprout later.
These disturbances may be exacerbated by fire, which may further damage stems, resulting in more sprouting or a temporary flush of nutrients to the system; ailanthus’ uptake of these nutrients may be faster than that of associated native species (Medina-Villar et al., 2016). These disturbances followed by an extended drought could increase mortality of young ailanthus sprouts and seedlings resulting from the openings.
Disturbances that result in stem damage are likely to initially result in abundant sprouting and suckering of any ailanthus damaged in the disturbed area along with new ailanthus seedlings if there is a nearby seed source. Immediate surveys and treatments of ailanthus (and other invasive species) after such a disturbance will increase the likelihood that managers can contain their spread.
Goods
Ailanthus is used for timber in its native range (Hu, 1979). Its wood is ring porous with wide, attractive white or yellowish white sapwood and yellow-gray or yellow-brown heartwood that is only slightly darker in color than the sapwood (Grosser, 1977). The wood is similar in density and appearance to some ash and hackberry (Celtis) species, and its working, finishing, machining, drying, and gluing properties are similar to ash (Brandner and Schickhofer, 2010; Flynn and Holder, 2001). The disagreeable odor that is present in the vegetative parts of the tree is not present in dry wood (Asaro et al., 2009; Flynn and Holder, 2001). Its wood specific gravity, strength, stiffness, and hardness are also comparable to ash and hackberry, and, like these species, the wood is used in construction, tool handles, furniture, and sporting goods (Mahmood, 2006; Panyotov et al., 2011) (see appendix).
Ailanthus’ relatively dense wood is susceptible to splitting, although processing with a dry heat or oil heat treatment or acetylation (or combination thereof) may improve ailanthus’ wood quality for veneer use (Miao et al., 2014; Németh et al., 2020). Its wood can be processed for paper, though its pulp fibers are short and wide, making them susceptible to tearing (Baptista et al., 2014; Panayotov et al., 2011).
Termites rarely infest ailanthus wood, making its mulch a potential deterrent against termites (Grace, 1997). However, it is not resistant to decay (Flynn and Holder, 2001). Ailanthus sawdust may be used as an absorbent material for acid yellow 29, a toxic compound used to dye fabrics and commonly found in wastewater (Rahman et al., 2021).
Ailanthus fuelwood properties are comparable to those of white oak, black walnut, and birch (not identified to species in article; Hu, 1979). The very low ash content and high calorific values of ailanthus trees that are greater than 20 years of age make it suitable for solid biofuel production, including pellets (Kamperidou et al., 2017; Terzopoulou and Kamperidou, 2022), and its ability to coppice after cutting makes it a sustainable source of fuel (National Research Council, 1980). In addition to its use as fuelwood, it is used for charcoal production (Fry, 2010; Shah et al., 2014).
Ailanthus was originally imported to the United States to be grown in urban environments because of its tolerance to pollution, poor soil, and aridity (Miller, 1990; National Research Council, 1980). It is used in honey production (Melville, 1944), as food for silk-producing caterpillars (Hu, 1979; Kowarik and Säumel, 2007), and as a source of essential oils (Sladonja et al., 2015). Leaves can be used in yellow dye production and are edible after boiling, but due to their bad smell and toxicity this was done only in times of food scarcity (Tanaka, 1976).
About 221 chemical compounds have been characterized from ailanthus’ dried bark. These support a range of pharmacological activities including antitumor, anti-inflammatory, and antiviral properties and the potential to treat cancer, female reproductive ailments, epilepsy, chronic diarrhea, and gastrointestinal inflammation (De Martino and De Feo, 2008; Li et al., 2021). At least five quassinoid compounds found in these bark extracts of ailanthus have antimalarial properties, including against the chloroquine-resistant Plasmodium falciparum (Bray et al., 1987). A sixth compound, 6α-tigloyloxychaparrinone, may be less toxic to infected cells than ailanthone and is thus a potentially safer antimalarial quassinoid (Okunade et al., 2003). Ailanthone has also shown potential for the treatment of cancer (Ding et al., 2020).
Leaf extracts from ailanthus have also proven to inhibit some bacteria of concern, including Listeria monocytogenes, Staphylococcus aureus, Bacillus subtilis, Escherichia coli, Salmonella typhimurium, and Pseudomonas aeruginosa, as well as food spoiling bacteria (Poljuha et al., 2017; Rahman et al., 2009). Leaf extracts are also effective against the fungus Candida albicans (Gürbüz and Kahramanoğlu, 2021; Poljuha et al., 2017). Extracts from the samaras have quassinoids that are active against the tobacco-mosaic virus, an easily transmitted virus that affects a wide range of plants, especially members of the tomato and potato family (Solanaceae) (Tan et al., 2020).
Ailanthus bark contains compounds that can act as herbicides, preventing germination and growth of weeds (Caser et al., 2020; Heisey and Heisey, 2003; Tsao et al., 2002). Ailanthus also contains essential oils (e.g., α-curcumene, α-gurjunene, ƴ-cadinene, α-humulene, β-caryophyllene, caryophyllene oxide, and germacrene D), that may act as insect repellents and insecticides. Leaf and bark extracts, particularly quassinoids, are potent fumigants (De Feo et al., 2009; Kozuharova et al., 2022). Ailanthus leaf extracts are also slightly insecticidal against yellow fever mosquito larvae (Aedes aegypti) (Tsao et al., 2002) and other mosquitoes (Aedes aegypti and Culex quinquefasciatus) that serve as vectors of disease (Wallace et al., 2021).
Ecosystem Services
Although ailanthus is associated with areas of high diversity in the early stages of its invasion of a site, it is documented to reduce local plant biodiversity (Brooks et al., 2021; Terzi et al., 2021), change soil invertebrate and microbe compositions (Motard et al., 2015), and prevent regeneration of some native species, including oaks (Quercus spp.), by having a negative allelopathic effect on their mycorrhizal associations (Bauman et al., 2013). However, ailanthus-invaded sites may indirectly contribute to the biodiversity of other trophic levels by serving as a food source, pollen source, or nesting site (Aldrich et al., 2008; Antonov et al., 2007; Smith et al., 2014; Thompson, 2008), partially replacing the function of any native species it may have displaced.
During its first 6 months of growth, ailanthus can significantly decrease the total nitrogen in the soil at rates higher than found with associated native species (Medina-Villar et al., 2016). However, ailanthus also produces abundant leaf litter which decomposes rapidly, releasing nitrogen at rates 1.7 to 2.2 times higher than found in uninvaded sites (Medina-Villar et al., 2015). Ailanthus as a rapidly growing tree could serve as a sink for carbon; its average carbon sequestration rates, at least in an urban environment, are lower than 7 other urban tree species (e.g., black locust and sycamore [Platanus]), but higher than 12 other measured species (Tuğluer et al., 2017).
Ailanthus’ presence may have a negative effect on recreation due to its smell, its pollen being an allergen (Mousavi et al., 2019), and the potential for dermatitis for some who encounter the foliage or bark (Feret, 1985; Whiticar and Harvey, 2009). Ailanthus' association with large populations of spotted lanternfly makes it a source of sticky honeydew covered in black mold. This honeydew or mold (or both) may then attract ants, bees, hornets, and wasps (Dara et al., 2015).
Ailanthus is utilized in erosion control, shelter belts, reforestation, nutrient cycling, and soil formation in areas where rapid growth is desired, such as in mined sites being reclaimed and other degraded areas (Enescu, 2014; Enescu, 2022). Ailanthus is still sold online as an ornamental tree that provides beauty, shade, and privacy belts.
Ailanthus is a very common tree in urban areas of the midwestern and northeastern United States (Hu, 1979). It is found in a variety of different urban habitats including public and private open and confined spaces, archaeological sites, railways, roadsides, sidewalk cracks, wall fissures, and manholes, but it proliferates in urban flower beds, gardens, parks, tree-lined streets, and urban green spaces (Casella and Vurro, 2013). There is a notable association of ailanthus with transportation networks connecting urban areas, suggesting a small-world architecture to its spread among urban areas (Aldrich et al., 2010; Aldrich et al., 2015; Huebner, 2003; Rozenblat and Melanҫon, 2013). Larger urban forest patches (95 to 126 ha [235 to 311 acres]) have significantly fewer nonnative tree species than small urban forest patches (0.05 to 0.65 ha [0.1 to 1.6 acres]) and ailanthus may be more likely to be found in the smaller urban forest patches (Doroski et al., 2022) and is an important component of urban vacant lots (Riley et al., 2018). Ailanthus is also documented in old-growth urban forest gaps (Massad et al., 2019). In Baltimore, Maryland, most of the ailanthus trees are found in a land use associated with a form of transportation, such as roads and railways, representing 58 percent of all trees found in this urban land use (Nowak et al., 2004). Given its abundance in urban areas and along transportation corridors, these environments may still serve as source populations of spread into other urban areas even with the elimination or reduction of intentional plantings of ailanthus. However, the genomic make-up of ailanthus in smaller cities (e.g., as found in West Virginia) manifests genetic drift and founder effects rather than intermixing, suggesting boundaries still exist in some areas that may limit long-distance dispersal and novel genetic structures may develop and persist (Aldrich et al., 2010). Management (i.e., active removal) of ailanthus in urban and natural areas may also help to inhibit long-distance gene flow and admixture (Brusa and Holzapfel, 2018).
Forests in Cities
Urban areas as heat islands may correspond to greater seed germination and survival of the somewhat-thermophilic ailanthus (Zaraś-Januszkiewicz et al., 2014). Seed dispersal along impervious, flat surfaces that characterize many urban sites may be extended by as much as 400 m (1,312 feet) (Kowarik and von der Lippe, 2011) with 100 m being a typical seed dispersal distance for ailanthus (Landenberger et al., 2007). Human activity associated with urban areas, such as cars, railways, transfer of fill and soil material may facilitate long-distance dispersal events (Kowarik and von der Lippe, 2011). Possibly because of higher seed predation rates found in urban forests than are found in rural communities (Bowers and Breland, 1996), urban forest seedbanks lack native seeds and are often dominated by nonnative species, including ailanthus (Pellissier et al., 2008).
Seedling root growth may differ due to the increased soil compaction found in urban areas, though urban forests may be less compacted than other urban areas. In noncompacted soil, ailanthus will form a long tap root with two to three coarse lateral roots. In contrast, in compacted soil, the tap root is significantly smaller and there is a proliferation of multiple secondary and tertiary branches coming off the lateral roots (fig. 7) (Pan and Bassuk, 1985). When compared to other common street trees, such as sweetgum (Liquidambar styraciflua) and Norway maple (Acer platanoides), ailanthus has just over three times the amount of lateral growth and two times the shoot growth in only 2 years (Pan and Bassuk, 1986).
Figure 7—Root system of a 2-year-old ailanthus sapling that was growing in compacted soil. USDA Forest Service photo by Cynthia D. Huebner.
Once established as a sapling, ailanthus has the potential to produce seed as early as 3 to 4 years of age (Bory and Clair-Maczulajtys, 1977). Ailanthus trees in urban areas tend to have a high rate of mortality (e.g., 17.6 percent in Baltimore), but this rate is highest in transportation zones, where it is 35.1 percent; ailanthus’ mortality rate is only 3.6 percent in residential areas and 0.0 percent in low density residential areas (Nowak et al., 2004), which are likely to be irrigated, mulched, and fertilized. Studies of some urban forests with ailanthus share a pattern of a lack of ailanthus saplings but gains in trees and seedlings over time (Piana et al., 2021; Templeton et al., 2019; Turner et al., 2007). However, in other urban forests (e.g., Chicago, Illinois and Syracuse, New York), most increases in ailanthus have been of smaller trees or saplings (Nowak et al., 2013; Nowak et al., 2016; Pan and Bassuk, 1986). These differences in mortality among size classes may reflect differences in canopy opening in each of these urban forests. Survivorship of ailanthus seedlings under closed canopy urban forests may be low, but saplings can survive as sprouts under closed-canopy conditions even in natural forested sites (Kowarik, 1995).
Trees in Planted Urban Landscapes
In the United States and Canada ailanthus is listed as a problem invasive and is on many city “do not plant” or similar street tree or landscaping species lists. However, its commonness in some urban residential areas in the United States as both saplings and trees suggests it was and may still be being planted by individual home dwellers (Todd, 2014). Some city planners in eastern Europe are considering inclusion of ailanthus in their plans for urban tree plantings, where they might strategically plant or maintain existing ailanthus in high pollution zones and other extreme conditions, where native species or noninvasive exotics are unable to survive (with large distances between male and female trees to prevent seed production). The premise is to provide some ecosystem services that are missing from these extreme areas (Pindaru and Nita, 2023).
Despite its known invasiveness, there are some perceived benefits of ailanthus in urban areas. Ailanthus’ rapid growth, tolerance of pollution and soil compaction, attractive form and seed clusters, ability to reduce stormwater runoff, potential to reduce energy costs by serving as a shade tree by alleviating the heat associated with urban areas, and provision of nest sites for some native birds (Smith et al., 2014) and pollen for pollinators (Aldrich et al., 2008; Thompson, 2008) where native plants are lacking may be beneficial in urban environments (Pindaru and Nita, 2023; Riley et al., 2018). An estimated economic value of these services (based only on costs associated with pollution and soil compaction and reductions in stormwater runoff and energy costs) have been estimated as a cost savings (U.S. dollars) of $2,931 ha-1 ($1,187 acre-1) in vacant lots, $1,320 ha-1 ($534 acre-1) in downtown (city-center) residential lots, and $861 ha-1 ($349 acre-1) in suburban residential lots (Riley et al., 2018).
Ailanthus also has features that are disservices in urban areas. It is difficult to remove once planted because of the persistent root and stump sprouts. These roots and sprouts also damage infrastructure, such as streets, sidewalks, powerlines, and pipes (Casella and Vurro, 2013; Sladonja et al., 2015). Its large leaves have stiff rachises that add debris to drains and gutters. Ailanthus is malodorous, the pollen is an allergen, and the sap from the bark and leaves may cause dermatitis (Feret, 1985). Finally, ailanthus growing in urban areas may also be contributing to the loss of several native species in these areas (Doroski et al., 2022; Motard et al., 2011) as well as a source of its spread into natural areas (Landenberger et al., 2009; von der Lippe et al., 2005).
No cost-benefit analysis of ailanthus in urban areas has been completed, but a survey revealed that laypeople and those over 60 years of age in Berlin, Germany see ailanthus as having a positive role in their city (Kowarik et al., 2021). Calculating and sharing information on the costs of native species displacement, damage to infrastructure, and negative health effects may result in more-informed perspectives about ailanthus from people of all age groups.
Ailanthus is a comparatively well-studied invasive tree among the trees invading North America, but there remain some gaps in research on the biology, ecology, and management of this species. These include:
- Ailanthus’ distribution and invasive status in Mexico need further clarification.
- Comparisons of how the different life-stages (e.g., sapling vs. tree) and origin (seedling vs. sprout) of ailanthus respond to different environmental conditions may help predict a population’s (with known age structure) response to different disturbances and changes in climate.
- The expanding range of ailanthus in its invasive range suggests it is adapting/evolving to novel environments. Genetic analyses are needed to evaluate the degree of adaptation that has occurred and that may occur in the future. This information may, in turn, give us more insight into predicting how ailanthus may respond to new environments (e.g., urban vs. natural areas), climate change as well as new pathogens and insects.
- The potential uses of a triploid genotype to control ailanthus needs further evaluation.
- Adapting to new situations also implies the potential to develop resistance to any proposed biocontrol agents, such as Verticillium nonalfalfae. Research is needed to determine if any resistance currently exists or at what rate it would develop.
- While it appears the allelopathic compounds of ailanthus break down quickly in most soils via soil microbes, there may be certain habitats or soil types with limited soil microbes that may make the compounds more detrimental to associated native plants, such as in the rare serpentine soil communities or in nearly soilless environments in urban areas. More research is needed to evaluate the differential impacts of these compounds across different habitat and soil types.
- Ailanthus’ allelopathic compounds could potentially be absorbed by fungi helping in the decomposition of stems and leaves. Would these fungi, if edible, then be potentially toxic if consumed or does the decomposition process break down their toxicity as is seen with other soil microbes?
- More research is needed in evaluating the potential benefits of ailanthus as a source of wood products (including timber), medicinal products, and important ecosystem services associated with urban (reduction of heat and stormwater runoff) and natural areas (source of food and habitat for native insects and birds) and how these may or may not outweigh the costs associated with its invasion, including the loss of biodiversity. It may be that these cost-benefit analyses are likely to vary with habitat types, with a gradient of high benefits to high costs coinciding with an urban to natural area gradient.
| Species | Specific gravitya | Modulus of rupture (kilopascals) | Modulus of elastic (megapascals) | Shear strength (kilopascals) | Side hardness (newtons) |
| Ailanthus (China)b | 0.67 | 81 000 | 10 500 | 11 700 | 4 600 |
| Ailanthus (Virginia)c | 0.62 | 97 000 | -- | 14 800 | 5 700 |
| Hackberryd | 0.53 | 76 000 | 8 200 | 11 000 | 3 900 |
| White ashd | 0.60 | 106 000 | 12 000 | 13 200 | 5 900 |
| Green ashd | 0.56 | 97 000 | 11 400 | 13 200 | 5 300 |
a Based on oven-dry weight and volume at 12 percent moisture content.
bSource for data on China-grown material: Cheng et al. (1992).
c Source for data on Virginia-grown material tested at Virginia Polytechnic Institute and State University, Blacksburg, Virginia: Asaro et al. (2009).
d Source: Forest Products Laboratory (2021).
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We thank 22 reviewers, who remained anonymous throughout the review process, and Cynthia F. Moser for their many comments and careful edits that much improved the final chapter. We thank James Miller, the author of the ailanthus summary in the preceding “Silvics of North America.” His contribution of including one of the few exotic species in the 1990 manual helped lay the foundation for inclusion of many more exotic tree species in the new silvics manual. We also thank Lorenzo Ferrari and Tessa Fenstermaker for their assistance with the literature search. Finally, we thank Cameron Stelly, Susan Iott, and Lauren Pile Knapp and the USNAP Exotic and Invasive Species Group for their guidance and keeping us organized and consistent.