a large kauri tree on New Zealand’s North Island; photo by F.T. Campbell
I applaud the growing attention to the impacts of deadly pathogens in forests. The focus is often on the resulting decrease in tree-based storage of atmospheric CO2. In an earlier blog I cited findings by Quirion et al. (2021) that forest diseases in the United States between 2001 and 2019 reduced mean forest carbon sequestration by 28%. Forests were converted from carbon storage to carbon sources. [Full citations for all sources provided at the end of this blog.] Earlier, Fei et al. (2019) had determined that compensatory growth by non-host species (spurred by reduced competition and greater resource availability after a dominant tree species dies) might partially offset CO2 decline, although only over long timescales. Chan et al. (2026) cite examples from redwoods in California to Dipterocarps in Southest Asia that document the importance of large trees in sequestering atmospheric CO2. See also Calders et al. re: Wytham Woods in the United Kingdom.
A pertinent current example is found in New Zealand. There, the oomycete Phytophthora agathidicida is killing the endemic kauri tree, Agathis australis. Kauri are among the largest and longest-lived trees on Earth. The average age of large individuals in old-growth stands is ~ 600 years; exceptional specimens reach 1,600 years. Typical diameters in old-growth stands are 1.3–1.5 m; exceptional specimens > 7 meters (Chan et al. 2026).
A year ago Simpkins et al. (2025) reported a modeling study that indicated that kauri dieback disease could cause a decrease in aboveground carbon stocks by 55% over 500 years. However, as Chan et al. (2026) point out, stand-level carbon sequestration and storage in a forest undergoing widespread mortality of the dominant tree species depend on growth and productivity of the replacement species. Even the magnitude of short-term dynamics depends on the relative abundance of the vulnerable host and the rate of its mortality and decline.
Therefore they explored the impact of kauri dieback in greater detail, including which species they thought likely to replace kauri and their carbon-sequestration capabilities at the stand level.
Although (regrettably) not the focus of their article, Chan et al. (2026) note that loss of the kauri tree also threatens biodiversity. Kauri strongly modify their environment through accumulation of a thick organic soil layer derived from litter. The layer is typically 30 to 50 centimeters deep but can reach 2 meters. These soils are acidic, drought-prone and nutrient-poor. As result the co-occurring plants comprise a unique suite of stress-tolerant plants. The vertical stratification of conifer and angiosperm crowns in kauri forests allows coexistence and unusually high total basal area of other plant species.
In these forests, kauri have dominated annual CO2 capture and long-term storage. Under pre-invasion conditions, Chan et al. (2026) this to continue. Previous work estimated that live vegetation biomass in kauri stands doubles when the stand moves from early- or mid-successional stages to mature. Three other species each contribute more than 10% of carbon storage: the tree fern Cyathea dealbata, and two small trees, Phyllocladus trichomanoides and Kunzea ericoides. All are early successional species that typically dominate in open or disturbed environments but decline as canopies close.
Chan et al. (2026) studied changes in live vegetation carbon stocks between two census periods ten years apart, 2011 – 2014 and 2021 – 2024. The five kauri-dominated plots included two plots each in early- and late-successional stages and one mid-successional stand. The study site was in the greater Auckland region of North Island. Phytophthora agathidicida is spreading through these forests, infecting kauri of all ages. Still, dieback intensity varies within the plots, so they were subdivided plots into 20 10 m × 10 m subplots.
Chan et al. (2026) tried to measure two distinct processes: the increase in carbon stocks coming from recruitment and growth of live stems (= productivity) and loss of sequestration and storage due to tree mortality. The two phenomena might respond differently to kauri dieback and initial carbon stocks. While increasing disease severity and increased mortality both result in lower net accumulation, it is important to disentangle the effects of increased loss, reduced productivity, or the combination.
kauri killed by dieback disease; 149 kauri dieback 02-868×1300
Findings
Carbon sequestration declined proportionately with increasing disease severity due to both reduced productivity and greater carbon loss from mortality. Because kauri contributed the largest share of carbon sequestration and storage due to the enormous size of some individuals, death of higher numbers of large kauri led to the largest reductions in these services. That is, the greatest observed decrease in carbon sequestration and storage was in those stands that initially had the highest levels of carbon storage. These were stands in mid- and late-successional subplots rather than in early-successional subplots. The higher the initial mean carbon stock level, the more pronounced the observed decline in sequestration.
This means that kauri dieback threatens both the carbon sequestration and long-term carbon storage capacity of kauri forests, particularly in mature stands comprising the largest trees.
Chan et al. (2026) conclude that the carbon stocks of these mature kauri stands will not stabilize until after substantial depletion of kauri biomass. As a result, long-term carbon storage capacity is markedly reduced.
Meanwhile, healthy kauri stands continued to sequester carbon as the trees grew. Chan et al. (2026) expect living kauri trees to continue adding to their basal area and other productive species can co-exist under the vertically stratified crowns.
In stands that had low carbon sequestration levels at the time of the first census (early successional stands), the stand maintained positive carbon sequestration until the disease reached a higher intensity. This was because some of the other tree species in the stand raised their proportional contributions to CO2 sequestration. Chan et al. (2026) concluded that these stands were likely stabilize at a reduced level of biomass, effectively lowering their long-term carbon storage and sequestration potential.
The steep decline in carbon sequestration and storage associated with higher initial carbon stocks reveal the disproportionate vulnerability of mature kauri stands. The kauri contribution to stand biomass is unlikely to be replaced by a structurally similar species, e.g., the long-lived, large conifer rimu (Dacrydium cupressinum), until passage of many years. Early-successional species are expected to colonize the damaged forests rapidly, but their lower wood densities and fast maturity/senescence confer limited carbon storage capacity.
a rimu tree; photo by F.T. Campbell
Besides, some of the species likely to replace (temporarily) dead kauri — Myrtaceae such as Kunzea ericoides — are susceptible to another introduced pathogen, myrtle rust (Austropuccinia psidii) (Lantham et al. 2025).
So kauri dieback not only reduces current carbon stocks but succession dynamics do not support future sequestration potential, leading to a long-term decline in the carbon storage capacity of these forests.
Chan et al. (2026) call for urgent efforts to prevent spread of the pathogen Phytophthora agathidicida and injection of phosphite to reduce disease symptoms. Both strategies should target mature kauri stands. Research on phosphite treatments is under way. They also suggest research on disease progression and associated changes in carbon stocks across multiple pools, including coarse and fine woody debris and deep organic soil layers.
Chan et al. (2026) report that the disproportionate contribution of large, long-lived individual trees to carbon storage is a worldwide phenomenon.
Extraneous thought: I am glad that Chan et al. (2026) cite examples of high-impact pathogens from the Southern Hemisphere, i.e., Phytophthora cinnamomi in Australia. While chestnut blight certainly had severe impacts, so have many others, e.g., white pine blister rustCronartium ribicola. Some arthropods have also proved damaging on several continents, including emerald ash borer (Agrilus plannipennis), or across widely separated island systems, e.g., Erythrinagall wasp (Quadrastichus erythrinae) and cycad scale (Aulacaspis yasumatsui). I regret that I can find no recent information on the scale.
SOURCES
Calders, K., H. Verbeeck, A. Burt, N. Origo, J. Nightingale, Y. Malhi, P. Wilkes, P. Raumonen, R.G.H. Bunce, M. Disney. Laser scanning reveals potential underestimation of biomass carbon in temperate forest. Ecol Solut Evid. 2022;3:e12197. wileyonlinelibrary.com/journal/eso3
Chan, P.J., T. Elliott, H.R. Lai, B. Burns, L. Schwendenmann. 2026. Consequences of a dieback disease on carbon stocks and fluxes in forests dominated by a susceptible foundation species. Plant Ecology (2026) 227:97 https://doi.org/10.1007/s11258-026-1669-4
Fei, S, R.S. Morin, C.M. Oswalt, and A.M. Liebhold. 2019. Biomass losses resulting from insect and disease invasions in US forests. Proceedings of the National Academy of Sciences. August 27, 2019 vol. 116 no. 35 www.pnas.org/cgi/doi/10.1073/pnas.1820601116
Latham, M.C., A. Lustig, N.M. Williams, A. McDonald, T. Patuawa, J. Chetham, S. Johnson, A. Carrington, W. Wood, and D.P. Anderson. 2025. Design of risk-based surveillance to demonstrate absence of Phytophthora agathidicida in NZ kauri forests. Biol. Invasions (2025) 27, no.26 https://doi.org/10.1007/s10530-024-03501-4
Quirion BR, Domke GM, Walters BF, Lovett GM, Fargione JE, Greenwood L, Serbesoff-King K, Randall JM and Fei S (2021) P&P Disturbances Correlate With Reduced Carbon Sequestration in Forests of the Contiguous US. Front. For. Glob. Change 4:716582. [Volume 4 Article 716582 doi: 10.3389/ffgc.2021.716582
Simpkins, C.E., P.J. Bellingham, K. Reihana, J.M.R. Brock, G.L.W. Perry. 2024. Evaluating the effects of two newly emerging plant pathogens on North Aotearoa-NZ forests using an individual-based model. Ecological Modelling, www.elsevier.com/locate/ecolmodel https://doi.org/10.1016/j.ecolmodel.2024.110938
Posted by Faith Campbell
We welcome comments that supplement or correct factual information, suggest new approaches, or promote thoughtful consideration. We post comments that disagree with us — but not those we judge to be not civil or inflammatory.
For a detailed discussion of the policies and practices that have allowed these pests to enter and spread – and that do not promote effective restoration strategies – review the Fading Forests report at http://treeimprovement.utk.edu/FadingForests.htm
old-growth forest in Tioresta Research Natural Area, Allegheny National Forest (Pennsylvania); photo by Nicholas T via Flickr
The implied or explicit recommendations are not new. But they are supported by new data!
Several measures of forest health drawn from survey plots of the Forest Inventory and Analysis (FIA) show that federally-owned forested lands – in both National forests and National parks – are healthier than neighboring forests under different ownership. Forest inventory plots in federal forests have significantly greater tree species and structural diversity and evenness; basal area and biomass per hectare; and seedling density. They are also less invaded by non-native plants. These findings hold across the continental United States, four regions (North, South, Rocky Mountains, and Pacific Coast) analyzed separately, and 22 common forest types.
The consistent finding of more “mature forest” features in eastern National parks and National forests – regardless of longer-term land-use histories (Potter et al. 2026b) – seems to me to indicate that recent and current management practices might overcome influences of centuries of pre-protection land use histories of exploitation and degradation.
1. Forest Health Inside vs. Outside Federal Forested Lands
Legally designated USFS National forests comprise ~19% of the Nation’s total forest area (~59 million out of 310 million hectares). Since adoption of the Forest and Rangeland Renewable Resources Act of 1974, the National Forest System (NFS) within USDA has been managed to maintain “appropriate forest cover . . . to secure the maximum benefits of multiple use sustained yield management in accordance with land management plans”. The Department of Interior’s National Park Service units are managed to conserve unimpaired the natural and cultural resources and values in the system for the enjoyment, education, and inspiration of this and future generations (National Park Service “Organic Act” of 1916).
Potter et al. 2026b assessed the ability of National forests to provide ecosystem services and forest products in the face of dramatically increasing threats: (1) more intense and frequent wildfires; (2) uncharacteristically long and hot drought conditions; (3) mortality and compositional changes caused by invasive forest pests; (4) competition from invasive plant species; and (5) forest fragmentation and conversion to other land uses. To do this, they asked whether National forests (1) have greater biodiversity – as measured by species and structural diversity; (2) are denser and encompass more biomass; (3) experience more regeneration; and (4) are less invaded by non-native plants. The FIA surveys were carried out from ~2010 to ~2022. As noted, they analyzed each comparison across the conterminous states, the four regions, and the 22 forest types.
Potter et al. (2026b) and an earlier study by Miller et al. (2016) agree that ownership determines forest management, specifically the level of anthropogenic disturbance allowed. Miller et al. demonstrated that, in the eastern United States, forests in National parks have greater live tree basal area, a larger volume of coarse woody debris, and greater proportions of late-successional tree species than nearby forests. They also have lower rates of tree growth and mortality.
Potter et al. (2026b) found that all biodiversity indicators except number of saplings per hectare were higher for plots inside the NFS than neighboring forests under other ownerships. This was true across the continent and in all four regions.
In some of the 22 forest types the indicators were particularly strong. Thus, basal area was notably higher on NFS forests populated by loblolly/shortleaf pine, Western larch, hemlock/Sitka spruce, maple/beech/birch, and oak/hickory forest types. Tree species richness was significantly higher on NFS plots for Western larch, elm/ash/cottonwood, longleaf/slash pine, pinyon/juniper, and loblolly/shortleaf pine. Differences in regeneration were seen most prominently in loblolly/shortleaf pine, oak/hickory, oak/pine, and white/red/jack pine. The authors do not speculate on the reasons.
a large loblolly pine in Congaree National Park; via Picaryl
The FIA data showed that invasive plant richness and cover were both greater on plots outside National forests across the Continent and in all four regions. However, invasive plant richness did not differ by ownership for some specific forest types: elm/ash/cottonwood, pinyon/juniper, Western larch, and hardwood woodlands. (National parks are even less invaded by non-native plants than National forests; see below.)
2. Regions Differ – Why?
Geographic regions differed noticeably. Tree species richness and evenness; diversity and evenness of tree height classes; biomass/ha; seedlings and saplings per hectare; and invasive plant diversity and cover were higher in the East – in both USFS Northern and Southern regions. Tree species richness was by far highest in the South – especially in the southern Appalachians and on the Cumberland Plateau. Seedling density was highest in the North, moderately high in the South. In the North, tree species and tree height evenness were both higher on non-federal lands.
Invasive plant richness was highest in some parts of the South. [See a more detailed discussion of invasive plants below.] Tree diameter class diversity and evenness were significantly higher in the West. Basal area was by far the highest in the northern part of the Pacific coast region (Washington, Oregon, and California.). Seedling density was moderately high in some National forests in the Rocky Mountains. Tree diameter evenness was higher outside National forests in the Rocky Mountains.
These differences had been expected given the marked regional dissimilarities in floristic, climatic, and edaphic factors, as well as ownership patterns.
Potter et al. (2026b) report that tree recruitment was stronger in the East than the Rocky Mountain and Pacific coast regions. I wish they had discussed how this finding relates to earlier findings (Potter and Riitters 2022) that numbers of species experiencing poor regeneration were highest in the Southeast, followed by the Northeast and Midwest. Miller et al. (2023) also found that tree regeneration was poor in National parks from Virginia to Maine. The latter study attributed this failure to a combination of browsing by overabundant deer and competition by invasive plants. The latter is presumably even more important in national forests, which are invaded by non-native plants at twice the rate of national parks. Nonfederal forests are invaded at even higher rates – up to 3 times greater. [See here for a study of improve regeneration in National parks which control their deer populations.]
Potter et al. (2026b) note that the higher forest health indicator values for NFS forests prevailed across the Continent despite notable differences in land-use and management histories and landscape contexts. Most NFS land in the West was put under federal management before arrival of European pioneers, so more of their groves are classified as old-growth or mature. In the East, nearly all lands incorporated into the National Forest System were privately owned until they were purchased in the 20th Century. At that time these forests were often highly degraded. Not only did the eastern National forests have to recover from overexploitation; they are still typically embedded in a matrix of other land ownerships, often largely cleared of trees.
In their analysis, Potter et al. (2026b) controlled for differences in site productivity and some environmental conditions e.g., elevation and slope, but not soil or hydrological conditions. So they believe that the suite of environmental factors do not explain the “older” attributes of eastern forests managed by the USDA Forest Service and USDI National Park Service,. These attributes generally include a more complex forest structure; higher tree species richness; and greater basal area and density of live trees (Miller et al. 2016).
What differs is management. Miller et al. (2016) say explicitly that protection from forest harvesting for many decades probably is an important explanation of why forests in the eastern National parks have greater tree species diversity than forests owned by other parties. Potter et al. (2026b) instead emphasize the impacts of different land-use histories outside of National forest boundaries. Much of this land (especially in the East) has experienced continuous, centuries-long agriculture and residential development and more acute disturbance pressures associated with higher human population density. They also note that managers of non-federal forests goals often call for harvests or other active management – which can reduce forest structure, species richness, size of live trees, presence of downed woody debris, etc.
I note that the consistent finding of more “mature forest” features in eastern National parks and National forests – regardless of longer-term land-use histories (Potter et al. 2026b) – indicates that recent and current management practices might overcome older influences. Many of these forests have enjoyed protective management for less than a century; the Forest and Rangeland Renewable Resources Act of 1974 was adopted just 50 years ago. I assert that the decisive factor might be the much lower presence – although not absence! – of deliberately planted non-native plant species on federal properties. [Again, see a more detailed discussion of invasive plant issues below.]
Potter et al. (2026b) also comment that regeneration on NFS land is almost exclusively natural. As discussed above, they consider this regeneration to be successful, in that seedlings generally are being produced in greater numbers than on surrounding forests in other ownerships. Again, I would like to learn how this finding relates to the earlier study by Potter and Riitters (2022).
Potter et al. (2026b) state that National forests benefit from their greater size and stability of management goals. These are necessary for large-scale approach to conserve or imitate the processes that create diversity
As Potter et al. (2026b) note, in the East, reserved areas like National forests (or to a greater extent, National parks) represent a small proportion of the landscape. Their distinctness from neighboring forestlands emphasizes their ecological value. They call for managers to be alert to potential invasions of non-native plants or tree-killing pests from the surrounding agricultural and developed land. In the West, private forests abut federal forested lands less often and thus have less influence on the status of forest health indicators on National forests.
3. Plant Invasions in Forests
Several studies by Kevin Potter, Kurt H. Riitters and colleagues have documented the extent of plant invasions in America’s forests. Data supporting these studies come from the same forest plot surveys conducted under the Forest Inventory and Analysis (FIA) program.
The region most invaded is Hawai`i: 83% of FIA plots have one or more invasive plant species. The region that ranks second is the East, specifically the 37 states comprising the USDA Forest Service’s former Northern and Southern regions. Nearly 53% of FIA plots in this region have one or more invasive plant species (Potter et al. 2024). FIA plots in the Rocky Mountain West had only 11% (Potter et al. 2026a).
student volunteers in Camp S.E.E.D. (Students Encouraging Environmental Recovery) program pulling invasive plants in Shenandoah National Park; via Picaryl
As noted above, invasive plants are particularly problematic in the South. There, invasive plants have been documented on 55.3 million hectares. In the Northern Region, an incomplete survey found invasive plants on 36.9 million ha. In some counties of the 37 states, 80% of inventoried forest plots contain invasive plants. Spread of these invaders is largely unchecked. The invasions’ extent and intensity are so great that their complete removal – or elimination of their impacts – is “practically impossible” (Potter et al., 2024; Potter et al. 2026a).
Plant invasions in eastern forests are undoubtedly worse than these data indicate because the records include only some of the non-native plant species present — those considered to be the worst invaders at the time regional lists were compiled (Potter et al. 2026a). Not included on the survey list are wavy-leaf basketgrass (Oplismenus undulatifolius), fig buttercup (Ficaria verna), shrub and creeping Euonymus, callery pear (Pyrus calleryana). Only one privet (Ligustrum vulgare) is included.
Furthermore, Bradley, Early and Sorte (2015) report that while non-native plant species are already more widely distributed than native species, the average invasive plant species inhabits only about 50% of its expected range.
Sites Most Heavily Invaded
Analyses of the FIA data reveal several overlapping definitions of which locations and situations have been invaded most severely. These factors are separate from the issue of ownership discussed above. Invasive plants are more numerous, widespread, and diverse in more disturbed or fragmented sites. One study found that 65% of plots that experienced greater fragmentation were invaded compared to 46% of plots that experienced low fragmentation. Invasion rates are highest in the “wildland-urban interface” (WUI). Unfortunately, the WUI is growing faster than any other land use type in the country – especially in the East (Potter et al. 2024). Because the East is so heavily developed, a plot’s distance from a road was nearly irrelevant. Furthermore, development in the WUI also promotes planting of non-native species that might then invade the nearby forest. See my more detailed discussion of these issues here.
Highly productive sites [defined as a site’s ability to grow industrial timber; Potter et al. 2026b] are also prone to invasion: 76% of highly productive plots were invaded compared to 40% of low-productivity plots. It is not known whether highly productive sites are inherently more invasible, or, instead, that such plots were converted to agriculture earlier, so exposed to human disturbance longer (Riitters et al. 2017).
A third analysis found that the best predictor of the odds that a site would be invaded was the site’s ecological province as defined by Robert G. Bailey in 1995 (Riitters et al. 2017)
Propagule Pressure: the Role of Deliberate Planting
Over 25 years (Reichard and White 2001) numerous scientists have documented the role of deliberate planting – especially ornamental horticulture – in facilitating introduction and spread of invasive plants. Kinlock et al. (2025) found that more than 1,600 plant species sold by nursery and seed catalogs over 200 years had “naturalized” somewhere in the continental 48 states. Fertakos and Bradley (2024) found that species were likely to establish if they were introduced to as few as eight locations. Beaury et al. (2023) found that half of 89 plant species recognized as invasive are sold in the same locations where they are invasive. Another 25 species are sold in an area that is currently unsuitable for those species, but that will become more suitable for invasion as temperatures warm.
Japanese barberry – invasive that is widely sold; photo by Matthew Beziat via Flickr
Both Potter et al (2024) and Potter et al. (2026a) note that the flora of suburban and rural residential landscapes is dominated by non-native plant species. I add that the people who live there promote plant invasions in various ways, including planting shrubs or flowers in the woods and dumping yard waste there. The older the human settlement, the more years for these plants to spread – assisted by birds, wind, or water. Go here for a more detailed discussion of these issues.
What Should We Do to Curtail Introduction and Spread of Invasive Plants?
a. Regulations The ornamental plant market – whether brick and mortar stores or internet sales — is interstate in scope. Regulations need to match (Beaury et al. 2023). This requires Congress to adopt a new federal law. Under the Constitution, the appropriate entity for regulating interstate commerce is the federal government. But the current statute (the Federal Noxious Weed Act) does not address long-established, widespread species.
States currently have the lead in regulating sales of horticultural plants. Beaury et al. (2023) and Evans et al. (2024) criticize state restrictions as outdated, limited to a few weeds that plague agriculture, and irregularly enforced. The result is a checkerboard of places where a species is offered for legal sale next to places where that sale is prohibited. Finally, the regulations are reactive; they rarely include plants in anticipation of their spread to new areas. Evans et al. (2024) urge state regulators to prioritize those species in the ornamental trade that are projected to remain or become abundant under evolving climate conditions.
b. Voluntary Actions
Potter et al. (2024) call for efforts to encourage homeowners to plant more native and environmentally friendly private landscapes. They concede the complication that some non-native – even invasive – species provide valued ecosystem and cultural services. They also suggest that local governments adopt land-use planning rules that protect forests of high conservation value. They do not discuss the extreme improbability of the latter action given the magnitude of predicted land-use changes in the country, powerful demographic factors driving them, and lobbying clout of affected economic interests, including the nursery industry.
Many citizen associations – native plant societies, regional or state invasive plant councils, etc. – are pursuing the education approach. (See the websites for state native plant societies, Southeast Exotic Pest Plant Council, Mid-Atlantic Invasive Plant Council, Midwest Invasive Plant Network, and Virginia Invasive Plant Coalition.) These voluntary efforts have yielded some success. But they have not resulted in adequate protection for our ecosystems.
c. Land-Managing Agencies
Many land-managing agencies work with local and regional groups to monitor and remove invasive plants, e.g., Blue Ridge PRISM. Databases that verify and post their findings, e.g., iNaturalist and EDDmapS can provide early warning of new invaders. See here for a discussion of these matters.
Blue Ridge PRISM removing English ivy; photo courtesy of Rowena Zimmerman, Director of Virginia Invasive Plant Coalition & Blue Ridge PRISM
Potter et al. (2026a) also suggest that the USDA Forest Service update the lists of invasive plants to be included in future FIA surveys. I agree. See above for examples of highly invasive species not now included.
Of course, the influence of plants on ecosystems is broader than invasive species. Dr. Douglas Tallamy has demonstrated that even non-invasive, non-native plants can disrupt food webs.
SOURCES
Beaury, E.M., J.M. Allen, A.E. Evans, M.E. Fertakos, W.G. Pfadenhauer, B.A. Bradley. 2023. Horticulture could facilitate invasive plant range infilling and range expansion with climate change. BioScience 2023 0 1-8 https://doi.org/10.1093/biosci/biad069
Bradley, B.A., R. Early and C. J. B. Sorte. 2015. Space to invade? Comparative range infilling and potential range of invasive and native plants. Global Ecology and Biogeography
Evans, A.E., C.S. Jarnevich, E.M. Beaury, P.S. Engelstad, N.B. Teich, J.M. LaRoe, B.A. Bradley. 2024. Shifting hotspots: Climate change projected to drive contractions and expansions of invasive plant abundance habitats. Diversity and Distributions 2024;30:4154
Fertakos, M.E. and B.A. Bradley. 2024. Propagule pressure from historic U.S. plant sales explains establishment but not invasion. Ecology Letters 2024;27:e14494 doi: 10.1111/ele.14494
Kinlock, N.L., D.W. Adams, W. Dawson, F. Essl, J. Kartesz, H. Kreft, M. Nishino, Jan Pergl, P. Pyšek, P. Weigelt and M. van Kleunen. 2025. Naturalization of ornamental plants in the United States depends on cultivation and historical land cover context. Ecography 2025: e07748 doi:10.1002/ecog.07748
Miller, K. M., F. W. Dieffenbach, J. P. Campbell, W. B. Cass, J. A. Comiskey, E. R. Matthews, B. J. McGill, B. R. Mitchell, S. J. Perles, S. Sanders, J. P. Schmit, S. Smith, and A. S. Weed. 2016. National parks in the eastern United States harbor important older forest structure compared with matrix forests. Ecosphere 7(7):e01404. 10.1002/ecs2.1404
Miller, K.M., S.J. Perles, J.P. Schmit, E.R. Matthews, M.R. Marshall. 2023. Overabundant deer and invasive plants drive widespread regeneration debt in eastern United States national parks. Ecological Applications. 2023;33:e2837. https://onlinelibrary.wiley.com/r/eap
Potter, K.M and Riitters, K. 2022. A National Multi-Scale Assessment of Regeneration Deficit as an Indicator of Potential Risk of Forest Genetic Variation Loss. Forests 2022, 13, 19. https://doi.org/10.3390/f13010019.
Potter, K.M., K.H. Riitters, B.V. Iannone III, Q. Guo and S. Fei. 2024. Forest plant invasions in the eastern United States: evidence of invasion debt in the wildland‑urban interface. Landsc Ecol (2024) 39:207 https://doi.org/10.1007/s10980-024-01985-y
Potter, K.M., B.V. Iannone III, K.H. Riitters, Q. Guo, K. Pandit, C.M. Oswalt. 2026a. US Forests are Increasingly Invaded by Problematic Non-Native Plants. Forest Ecology and Management 599 (2026) 123281
Potter, K.M., Q. Guo, F.H. Koch, S. Lim-Hing, E.R. Matthews, and K. Pandit. 2026b. U.S. National Forests Are More Diverse, Denser and Less Invaded than Neighboring Forests. Forests 2026 17
Reichard, S.H. and P. White. 2001. Horticulture as a Pathway of Invasive Plant Introductions in the United States. BioScience 103. Vol. 51 No. 2. February 2021.
Riitters, K., K. Potter, B.V. Iannone III, C. Oswalt, S. Fei, Q. Guo. 2017. Landscape correlates of forest plant invasions: A high-resolution analysis across the eastern United States. Diversity and Distributions. DOI: 10.1111/ddi.12680
Tallamy, D.W. and K.J. Shropshire. 2009. Ranking Lepidopteran Use of Native Versus Introduced Plants
Conservation Biology, Volume 23, No. 4, 941–947 2009 Society for Conservation Biology
DOI: 10.1111/j.1523-1739.2009.01202.x
Posted by Faith Campbell
We welcome comments that supplement or correct factual information, suggest new approaches, or promote thoughtful consideration. We post comments that disagree with us — but not those we judge to be not civil or inflammatory.
For a detailed discussion of the policies and practices that have allowed these pests to enter and spread – and that do not promote effective restoration strategies – review the Fading Forests report at http://treeimprovement.utk.edu/FadingForests.htm
Cibotium glaucum in Hawaii Volcanoes National Park; F.T. Campbell
A year ago I blogged about the status of Hawaiian species in the face of high levels of bioinvasions. I now update that discussion as it pertains to one of the invaders, the Australian tree fern Sphaeropteris cooperi.
S. cooperi has successfully invaded multiple regions, including the southern coast of South Africa, the Azores and Mascarene Archipelagos, and Hawai`i. The invasion’s impacts have been assessed only in Hawai`i. Studies found that the non-native tree ferns have outcompeted the dominant native tree fern Cibotium glaucum (Cibotiaceae), and altered leaf litter composition hence soil nutrient cycling. This in turn affects species assemblages (van den Berg et al. 2025).
Chau, Walker and Mehltreter (2013) documented the litter and soil chemistry in Hawaiian rainforests. They found that S. cooperi produces more leaves that grow faster, contain more nitrogen and phosphsorus, and decompose faster than the leaves of the dominant native tree fern C. glaucum. They compared the effect of leaf litter from the native and tree ferns on the growth and nutrient content of four native angiosperm species when they were grown in N-rich forest soil and P-rich landslide soil. The results suggested that nitrogen availability is the strongest driver of growth. The introduced tree fern, S. cooperi, can thus prompt more rapid growth of some native HI plants. They point out, however, that under natural conditions, native plants must compete for these additional nutrients with various non-native plants, including S. cooperi. The ultimate impact, then, remains unclear.
van den Berg et al. (2025) report that other invaded sites are at great risk Both the Garden Route region of South Africa and La Reunion Island in the Mascarene Archipelago in the eastern Indian Ocean have high bioidiversity. Both contain large, globally-recognized protected areas established to protect the native biodiversity: UNESCO Garden Route Biosphere Reserve and Reunion National Park. Despite their biological importance, both are among the top10 most invaded countries/territories globally (Tuberlin et al. 2017).
Sphaeropteris cooperi; photo via Easyscape
While climate change is expected to reduce the extent of suitable habitat for the Australian tree fern in both South Africa and La Reunion, the current situation is troubling. At present the tree fern occupies a narrower range of climatic conditions in both the Garden Route and especially on La Reunion than in Australia. (On La Reunion, 13.74 % of the fern’s apparent niche remains unoccupied.) van den Berg et al. (2025) are not sure what factors might be limiting the tree fern’s spread. They do urge educational campaigns to persuade people living near the Biodiversity Reserve to avoid planting the non-native tree fern.
The tree fern’s actual niche is somewhat uncertain because predation by deer (non-native to Australia) might have reduced its reproduction. The authors mention this but do not speculate further on the possible response of the fern to the absence of such stresses on La Reunion.
SOURCES
Chau, M.M., Walker, L.R. and Mehltreter, K. An invasive tree fern alters soil and plant nutrient dynamics in Hawaii. Biol Invasions15, 355–370 (2013). https://doi.org/10.1007/s10530-012-0291-0
Turbelin,A.J., Malamud,B.D., Francis,R.A. 2017. Mapping the global state of invasive alien species: patterns of invasion and policy responses. Glob. Ecol. Biogeogr. 26,78–92. https://doi.org/10.1111/geb.12517.
van den Berg, M.L., G. Singh, E.J. McCulloch-Jones, M. Rouget, D.M. Richardson, T.B. Robinson. 2025. The invaded range of the tree fern Sphaeropteris cooperi is predicted to shrink in two southern hemisphere biodiversity hotspots. African Journal of Botany 178 (2025) 390-399
Posted by Faith Campbell
We welcome comments that supplement or correct factual information, suggest new approaches, or promote thoughtful consideration. We post comments that disagree with us — but not those we judge to be not civil or inflammatory.
For a detailed discussion of the policies and practices that have allowed these pests to enter and spread – and that do not promote effective restoration strategies – review the Fading Forests report at http://treeimprovement.utk.edu/FadingForests.htm
eastern (Canadian) hemlocks in Nova Scotia killed by hemlock woolly adelgid; photo by Celia Boone, NSDLF
Two recent studies show that climate change is driving changes to pest ranges. These pose a serious threat to forests of eastern Canada.
A study by Aubin et al. (2026; full citation at end of blog) assessed the risk that climate change would exacerbate the impacts of 14 non-native forest insect pests The 14 insect species were selected from an original group of 76 species using the following criteria:
1) They have the potential to cause mature tree mortality to at least one of the 37 tree species most abundant in Canada; and
2) They have been detected in Canada recently or have established populations actively spreading through the Canadian landscape.
Major Findings
Twenty-four tree species are affected currently by at least one of the 14 non-native insect pests (excluding Asian longhorned beetle; see below). (I list the tree and insect species below.) Four of them are already considered globally endangered or threatened due to invasive pests: black and white ash (Fraxinus nigra and F. americana), whitebark pine (Pinus albicaulis) and eastern hemlock (Tsuga canadensis). Another ash species (blue ash; Fraxinus quadrangulata) and another pine (limber pine; Pinus flexilis) are considered rare or threatened species in Canada.
brown spruce longhorned beetle; photo by Udo Schmidt via WikiMedia
Of the 24 species at risk, black spruce (Picea mariana) is most exposed because it grows in areas where eastern spruce gall adelgid and brown spruce longhorned beetle (Tetropium fuscum) are established. Expansion of these two pests could potentially reach 75% of black spruce biomass in Canada. However, Aubin et al. (2026) expect tree mortality to be limited because these insects target trees that are already stressed or weakened. Of course, the changing climate might increase the trees’ susceptibility. Thirty percent of the black spruce’s range is projected to be outside its current climatic niche by 2040.
The 24 tree species currently affected by at least one of the 14 non-native insects (excluding ALB) collectively constitute 3.2 billion tons of tree biomass. The impact is projected to increase more than four-fold — to 13.6 B tons of biomass — within two decades. Reinvasion by the Asian longhorned beetle would put at risk six additional tree species in the genera Acer, Betula, and Populus. Their jeopardy would add another 3.1 billion tons of live tree biomass to the “at risk” category.
Aubin et al. (2026) note that affected trees might remain alive but moribund, that is, lose their functional role within the ecosystem. for several years before finally dying. Therefore their analysis might underestimate pests’ impact on the forest. Their example is American beech – as weakened by beech bark disease. A reminder: beech bark disease causes widespread death of mature beech – opening the canopy and eliminating such wildlife-supporting aspects as nesting cavities and abundant crops of nuts. Vigorous root sprouting results in dense stands of young beech, crowding out other species.
Hotspots of greatest vulnerability
The analysis identified two hotspots of greatest vulnerability: north-central British Columbia near the border with Alberta, and along the St. Lawrence Seaway near the border with the United States. The pest pressures differ.
The threat to the British Columbia hotspot comes from expansion of mortality in dense pine forests caused by the native mountain pine beetle (Dendroctonus ponderosae). Aubin et al. (2026) describe a two-pronged impact from the changing climate: trees lose vigor because they become maladapted to the new growing conditions (temperature and drought); while the beetle increases the frequency of outbreak due to reduced overwintering mortality.
Along the St. Lawrence Seaway (southern Ontario and Quebec) the threat comes from a suite of non-native insects, including emerald ash borer, hemlock woolly adelgid, and beech scale. While emerald ash borer and beech scale were introduced directly to Canada by international trade, hemlock woolly adelgid spread across the border from the U.S. Further northward expansion of all three is projected under both low and high emission climate change scenarios.
beech bark disease in New Hampshire; photo by Eli Sagor via Flickr
Most alarming is that some regions in eastern Canada are vulnerable to invasion by all 14 insect species. Two additional pests loom: Aubin et al. (2026) fear northward expansion might reintroduce the Asian longhorned beetle or introduce the southern pine beetle (Dendroctonus frontalis). The latter has been expanding northward in the U.S.
Pests often move across the Canada-U.S. border. In addition to the five pests mentioned above, spongy moth, and two pathogens, the beech leaf disease nematode, and the oak wilt fungus, have spread from the U.S. into neighboring parts of Canada. The woodwasp Sirex noctilliowas probably introduced simultaneously to both countries. Winter moth and beech bark disease spread from Canada to the U.S. I worry that the brown spruce longhorned beetle might do the same.
Asian longhorned beetle: can this disaster be averted?
As noted above, the Canadians are alarmed by the prospect that the Asian longhorned beetle might be reintroduced – either by spread from extant populations in the United States or directly on imports from China. Aubin et al. (2026) note that the two earlier – successful! — eradication programs were expensive, costing an estimated CND$35.5 million. This expenditure is dwarfed by the costs estimated to arise from an unmanaged invasion: CDN$431 million annually in timber products and CDN$358 million annually in edible maple products. There would also be enormous ecological impacts, including threats to an additional 3,08 metric tonnes of tree biomass comprising ~24.96 tons of CO2 equivalent.
Canada’s central boreal forest is at lower risk both in terms of exposed tree biomass and number of invasive insect species present. The reasons are not understood. Aubin et al. (2026) suggest that the boreal ecosystem is more resistant to invasion due to a combination of environmental barriers and native natural enemies. For example, the introduced woodwasp Sirex noctilio did not cause widespread pine mortality in the region, probably due to antagonistic interactions with other subcortical species.
Although it does not fit this definition, Aubin et al. (2026) also evaluated the Asian longhorned beetle (ALB Anoplophora glabripennis), because of its huge impact if it is reintroduced to Canada (see above).
One of the focal groups, beech scale, Cryptococcus fagisuga, differs from the others because it is a vector of a tree-killing fungal pathogen (Neonectria coccinea); the scale itself does not cause notable harm.
Aubin et al. (2026) acknowledge that additional species represent a possible threat to Canadian forests. Therefore their study does not represent the total risk posed by all potential invasive insects in Canada, but provides a snapshot of selected, current vulnerabilities.
The 37 most abundant tree species in Canada together represent 88% of total mature forest tree biomass in Canada. They include 17 conifers and 7 deciduous trees:
Abies amabilis, Ab. balsamea, Ab. lasiocarpa; Acer rubrum, Ac. saccharinum, Ac. saccharum; Alnus rubra; Betula alleghaniensis, B. papyrifera; Callitropsis nootkatensis; Fagus grandifolia; Fraxinus american, F. nigra; Larix laricina, L. occidentalis; Picea engelmanii, P. glauca, P. mariana, P. rubens, P. sitchensis; Pinus albicaulis, P. banksiana, P. contorta, P. ponderosa, P. resinosa, P. strobus; Populus balsamifera, P. grandidentata, P. tremuloides; Pseudotsuga menziesii; Quercus rubra; Thuja occidentalis, T. plicata; Tilia Americana; Tsuga canadensis, T. heterophylla, T. mertensiana
The 14 insects collectively have 63 host tree species in Canada. The pine genus is susceptible to the largest number of pests. Genera found to be not vulnerable to any of the 14 insects are Acer, Callitropsis, Populus, Pseudotsuga, Tilia, and Thuja. The proportion of total exposed tree biomass in Canada varied by species, from 8% of Jack pine (Pinus banksiana) to 95% for red spruce (Picea rubens).
red spruce in West Virginia; photo by Famartin via WikiMedia
Although most of the 14 insect species are projected to benefit from larger areas of suitable climate in Canada over the next 20 years, there are interesting exceptions: European oak borer (Agrilus sulcicollis), brown spruce longhorned beetle, and winter moth (Operophtera brumata). Distributions of the host tree species are projected to change insignificantly over the 20 years covered by the study.
With northward expansion of suitable climates for 12 of the 14 species, large areas of the boreal forest will be exposed to potential invasion. The entire Canadian distribution of 13 of the 37 dominant tree species might be at risk: three Abies, two Betula, both Fraxinus, Picea mariana, Pinus contorta, Picea glauca, Pinus banksiana. Aubin et al. (2026) mention specifically Engelmann spruce (Picea engelmanii) and American beech (Fagus grandifolia).
Other Factors
The analysis did not consider possible alterations of the insects’ life history traits other than potential expansion of their distributions. Warmer temperatures can cause changes in voltinism, diapause periods, development rates, reproduction, and population growth; cumulatively, these changes might alter their invasion dynamics. The mountain pine beetle is an example. Populations have experienced outbreaks more frequently, so increasing the species’ invasion threat and severity. On the other hand, shifts in temperature and precipitation could decouple the phenology of trees and their associated pests, reducing insect survival. Finally, complex changes in tree tissue and their secondary defensive metabolites (see below) also could alter interactions between non-native insects and their new hosts – possibly exacerbating or mitigating the herbivores’ impacts.
Aubin et al. (2026) remind us that loss of a dominant species might lead to reorganization of forest composition and structure. They expect the impacts to be particularly critical in stands with low tree diversity, such as the pure jack pine stands in the eastern boreal forest. Loss of a foundation species might also profoundly disrupt ecosystem functions, carbon budgets, wildlife habitats and stand productivity. They cite cascading effect on aquatic invertebrate communities and invasions by non-native plant species following widespread death of ash trees caused by the emerald ash borer. Finally, death of some species reduces functional redundancy within tree communities, and a shrinking pool of viable native replacement species. The widespread planting of ash trees in urban areas after the demise of most elms is such a case.
American elms in Jeanne d’arc Park, Quebec; photo by Cephus via Wikimedia
The possibility that the tree hosts might increase production of defensive metabolites was corroborated by Mike Aucott in a different context. Dr. Aucott is retired from the New Jersey Department of Environmental Protection. He authored a guest blog in December 2022, in which he discussed changes in plant chemistry brought on by the 50% increase in atmospheric CO2 levels over the last century. By happenstance, Dr. Aucott engaged in an exchange of letters in Science (2/26/26), in which he reiterated the likelihood that plants, “fertilized” by access to this nutrient, might be better able to fend off insect attacks. (See the “Sources” section for references to additional information on this phenomenon.)
Compounding Threat: The Spruce Budworm in Eastern Canada
The spruce trees of eastern Canada face another pest threat: the native spruce budworm (SBW, Choristoneura fumiferana). It is already a major defoliator in North American boreal forests.
Boulanger et al. (2025; full reference at end of this blog) documented pronounced changes in the moth’s range, especially in the East. Over the past 60 years, suitable climate conditions for the budworm have expanded northward. On the other hand, winter mortality has increased in southern parts of its range due to warmer temps. Overall, the total area highly suitable for population growth remained virtually the same. Still, the budworm’s earlier activation might exacerbate its impact on the previously less vulnerable black spruce, Picea mariana. If so, this might fuel further increased population growth rates northward.
black spruce; photo by Laval University via WikiMedia
Like Aubin et al. (2026), Boulanger et al. (2025) found that Canadian forests in the east and Atlantic regions are likely to experience greater impacts on tree growth than are forests in western and central regions. The host most vulnerable to SBW, balsam fir (Abies balsamea), is a dominant or codominant species in the East and Atlantic regions. The fir is sparsely distributed in those areas of central and western Canada where the climate is becoming highly suitable for the insect. The frequent wildfires promote growth of young pioneer tree species, e.g., jack pine and trembling aspen, that do not support SBW.
At the most general level Boulanger et al. (2025) suggest that climate change might have already surpassed impacts of land use change on spruce budworm dynamics. Climate change puts additive and synergistic pressures on insects, which are already more sensitive than trees to climatic factors and able to adapt more quickly. As a result, climate change is becoming the most significant driver of recent declines in insect abundance and shifts in community structure, development, dispersal patterns, and phenology. (Again, see Aucott, above, for an alternative explanation.)
Boulanger et al. (2025) mention but do not discuss possible impacts of climate change and a shift in SBW distribution and tree hosts on a third trophic level, i.e., natural enemies. They note that many factors – not just climate suitability — influence trophic interactions. Another complication is that most SBW parasitoids require alternate hosts.
Boulanger et al. (2025) join others in urging forest managers to quickly adapt their management strategies to the novel climate-induced threats. They call for a proactive and integrated approach in forest management. Ecological research will be crucial to mitigate the compounded effects of climate change and to preserve the integrity & sustainability of forest ecosystems.
SOURCES
Aubin, I., A. Roe, B. Marquis, L. Boisvert-Marsh, J. Pedlar,S. Erni, B. Hamel, G. Lawrence, D. McKenney, T. Scarr. 2026. Vulnerability of Canadian forests to invasive insects under climate change. Accepted by the Canadian Journal of Forest Research.
Boulanger, Y., A. Desaint, V. Martel, M. Marchand, S. Massoda Tonye, R. Saint-Amant, et al. (2025) Recent climate change strongly impacted the population dynamic of a North American insect pest species. PLOS Clim 4(2): e0000488. https://doi.org/10.1371/journal. pclm.0000488
See also an article describing declining nutrient value of food crops in response to increased atmospheric C02 levels: https://www.washingtonpost.com/climate-environment/interactive/2026/carbon-pollution-diluting-key-nutrients-food/
Posted by Faith Campbell
We welcome comments that supplement or correct factual information, suggest new approaches, or promote thoughtful consideration. We post comments that disagree with us — but not those we judge to be not civil or inflammatory.
For a detailed discussion of the policies and practices that have allowed these pests to enter and spread – and that do not promote effective restoration strategies – review the Fading Forests report at http://treeimprovement.utk.edu/FadingForests.htm
SOD-infected rhododendron in a nursery; photo by Jennifer Parke, ODF
A group of scientists (See Khusnitdinova et al., 2026; full reference at the end of this blog.) contend that landscape interfaces—e.g., crop–forest edges, riparian zones, abandoned agricultural fields and orchards, and nursery–wildland transitions—are activezones of pathogen exchange. Biological and abiotic vectors collectively move pathogens from crops to wild plants, and vice versa. These exchanges create conditions speed up the evolution of pathogen aggressiveness and dispersal traits and promote the selection of generalist pathogen lineages capable of infecting both cultivated and wild hosts. In this way, crop-natural ecotones become not just passive transition zones but centers of adaptation.
The stronger or novel pathogens don’t stay in the specific local area; they are spread by a variety of human activities. Establishing large monocultures of crops and simplifying biological diversity at the landscape level boost inoculum production, limit host genetic diversity, and diminish natural regulation. Pathogens present in irrigation water can be spread during floods. Improperly sanitized green waste and compost can harbor viable oomycete propagules. Foot traffic and heavy equipment can move contaminated soil. Movement of infested plants for planting can transport the disease to a different continent. One example cited by Khusnitdinova et al. (2026) is the spread of numerous Phytophthora spp. from nurseries to forests and shrublands. A second example is rapid ʻōhiʻa death. They say it demonstrates that 1) a combination of human movement, forestry activities, and animal vectors can enable rapid local and landscape-scale spread; and 2) management measures (biobarriers, access control, restriction of animal movements, and phytosanitary inspection of planting material) can curtail that spread.
Meanwhile, the changing climate is causing shifts in the latitudinal and elevational distribution of plants and their associates; changing reproduction rates and latent periods; altering ranges and connectivity; and affecting disease incidence and severity. The direction is not always predictable; while drought or heat might reduce fungal and oomycete epidemics, the same conditions increase host stress and so might worsen disease outcomes.
Plant health scientists can use these concentrated geographic areas to focus plant disease surveillance. By integrating molecular and genomic tools with remote sensing and Geographic Information System (GIS)-based monitoring, plant health agencies can more quickly detect newly emerging diseases and implement effective action to counter the threat.
However, Khusnitdinova et al. (2026) warn that surveillance employing these technological advances can reduce the risk that a pathogen will “spill over” from an anthropogenic to a natural ecosystem or vice versaonly if pertinent sectors are transformed. Yes, they need resources: funding, staff, facilities. Also required is unification – or at least coordination. Khusnitdinova et al. (2026) advocate abandoning the compartmentalization that currently separatesforest health studies from invasive-plant and infectious-disease ecology studies. Instead, agencies should consider managed and natural systems together. They should conduct joint surveillance programs, share data standards, and coordinate management of the transition zones. In other words, apply a “One Health” landscape-based approach to the entire landscape.
Khusnitdinova et al. (2026) add that implementing such combined surveillance programs is especially vital in biodiversity-rich regions which have limited monitoring capacity. Might I suggest Hawai’i?
ohia trees killed by ROD; photo by J.B. Friday, UH
Other facts that challenge traditional phytosanitary practices
Khusnitdinova et al. (2026) provide strong evidence that pathogens change – sometimes quickly. Is the current regulatory system sufficiently flexible and agile to effectively address these developments?
First, pathogens’ host range is not fixed. Instead, it is a trait that changes quickly under the influence of alterations in effector repertoires, plant immunity genes, and environmental conditions (including those driven by human actions). Even small genetic changes—such as mutations, gene losses or gains, or horizontal gene transfers—can enable pathogens to infect new hosts or weaken previous infection barriers. They suggest that plant pathogens with broad host ranges, e.g., Phytophthora cinnamomi, can easily move between hosts in agricultural plantings, ornamental landscapes, and semi-natural vegetation within a relatively small region. Such frequent spillovers maintain inoculum in landscape mosaics and complicating eradication or containment efforts.
Khusnitdinova et al. (2026) note that host-range expansions have especially long-term consequence in forest ecosystems, where loss of a single tree species can change understory makeup, light and moisture patterns, related fungi and invertebrate communities, and ultimately, landscape diversity and function. They cite chestnut blight and sudden oak death in North America and ash dieback in Europe as examples.
In addition, Khusnitdinova et al. (2026) maintain that genetic recombination is now recognized as a fundamental driver of innovation in plant pathogen populations. Table 2 of their publication lists pathogens exhibiting well-documented and experimentally confirmed cases of recombination, hybridization, or other forms of genome exchange. Forest-related examples include several Phytophthora hybrids and the ash decline fungus, Hymenoscyphus fraxineus.
Phytophthora dieback in Western Australia
Khusnitdinova et al. (2026) add their voices to a growing chorus decrying a global forest health crisis. They say that repeated pathogen introductions—often via trade in plants and wood—have shifted many temperate and boreal forests into states characterized by higher tree mortality, increased dominance of opportunistic or disturbance-adapted species, and reduced functional diversity. These changes lead to reduced resistance [defined as the capacity to limit damage during a new outbreak] and resilience [defined as the speed and trajectory of post-disturbance regeneration and ecosystem reorganization]. They note that increasing tree species diversity is one of the few management interventions that succeeds in strengthening both forest resistance and resilience to pathogens—by decreasing host density for specialist pathogens and reducing continuous “fuel” for epidemics.
One step toward improving scientific understanding on the scale they advocate, in their view, is the European Holistic Management of Emerging Forest Pests and Diseases (HOMED) effort. HOMED combines plant pathology, forest ecology, and biosecurity. The emphasis is on early detection, risk assessment, and management of human-mediated pathways, incl plant trade and nursery systems. The initiative aims to limit pathogen establishment and spread while strengthening forest resistance and resilience under global change. Participants also try to provide practical solutions for stakeholders to manage emerging native and non-native pests and pathogens threatening European trees not only in forests, but also in nurseries, urban and rural areas.
USDA Secretary Brooke Rollins
I am inspired by the proposals in Khusnitdinova et al. (2026). In hopes that USDA will explore how to implement them, I presented a poster presentation at the annual USDA Research Forum on Invasive Species. In that poster I suggested that these ideas complement USDA Secretary Rollins’ Memorandum on departmental research priorities. The need for research to clarify scientific puzzles is particularly acute regarding tree-killing pathogens nematodes, etc.
I suggested prioritizing research on the following issues:
Setting up intensive monitoring programs targetting the agriculture/natural system interfaces, as recommended by Khusnitdinova et al. (2025). These authors describe useful technologies in molecular diagnostics, genomic surveillance, environmental DNA, and remote sensing to detect fungi, oomycetes, rusts, bacteria, and viruses. Kantor et al. (2025) define techniques applicable for nematodes.
Rapid analysis of potentially invasive species and their pathways of entry revealed by “early warning” systems [e.g., APHIS’ “PestLens” website; “door knocker” introductions; academic studies; and “unimportant” species introduced to the U.S. (e.g., Leptosillia pistaciae in California)].
Exploring ways (in addition to those suggested by Khusnitdinova et al. 2025) to shorten the time lag between introduction of a pathogen and its detection.
Incorporating into risk analyses information from sentinel garden program. Fund expansion of data collection and analysis to address asymptomatic plants, sampling techniques, and seasonality, as outlined by Drs. Eliana Torres Bedoya and Enrico Bonello (at the 2025 USDA Research Forum) and Raffa et al. (2023).
Over a somewhat longer-term, I suggested that research address these topics:
Find techniques to speed up determination of disease causal agents – which often remain obscure for years or decades. The International Plant Protection Convention (IPPC) link requires countries to name the causal agent before regulating disease hosts and vectors.
Determine which components of a “systems approach” are most effective against each type of pathogen – fungi, oomycetes, rusts, bacteria, viruses, nematodes, etc.
With state counterparts, explore ways to better curtail domestic spread of organisms once they have established in the United States.
Integrate socio-economic drivers of pest introductions into studies. E.g., why do some organisms suddenly spread to numerous countries over a period of a few years?
Greatly expand efforts (in house and by collaborators) to breed trees resistant to established and newly detected pathogens.
Increase research supporting biocontrol.
As I have frequently complained in the past, the international phytosanitary system has failed to protect Earth’s forests and other natural ecosystems from non-native plant pests (or invasive plants). This failure has been documented by Weed, Ayres, and Hicke (2013), Fei et al. (2019), Quirion et al. (2021) for North America; and Gougherty (2023), Wu (2023), Sitzia et al. (2021), Martinac et al. (2025) and Khusnitdinova et al. (2025) from a global perspective.
Challenges:
Most microorganisms are unknown to science – “unknown unknowns”.
Scientists usually cannot predict the impact of known micro-organisms on new hosts under novel environmental conditions.
The World Trade Organization’s SPS Agreement and the International Plant Protection Convention (IPPC) demand unachievable levels of specificity re: a potential pest’s impact.
Most tree-killing pathogens are detected after they have entered the forest.
Agencies assign a low priority to protecting natural ecosystems from bioinvasion.
Resources (funds, staffing, etc.) are unreliable for agencies carrying out the full range of efforts, from assessing various risks to restoring pest-resistant trees to the forest.
SOURCES
Fei, S., R.S. Morin, C.M. Oswalt, & A.M. 2019. Biomass losses resulting from insect & disease invasions in United States forests
Gougherty, A.V. (2023) Emerging tree diseases are accumulating rapidly in the native & non-native ranges of Holarctic trees. NeoBiota 87: 143–160. https://doi.org/10.3897/neobiota.87.103525
Kantor, C., Teixeira, M., Kantor, M., and Gleason, C. 2025. Tiny Invaders, Big Trouble: Emerging Nematode Threats in the United States. Phytopathology 2025 115:587-595 https://doi.org/10.1094/PHYTO-09.-24-0290-IA
Khusnitdinova, M., V. Kostyukov, G. Nizamdinova, A. Pozharskiy, Y. Kydyrbayev and D. Gritsenko. 2026. Cross-Ecosystem Transmission of Pathogens from Crops to Natural Vegetation. Forests 2026, 17, 76
Martinac, M-L., F. Ningre, A. Dowkiw, N.Le Goff, B. Marcais. 2025. High host density favour ash dieback Preprint Plant Pathology
Quirion BR, Domke GM, Walters BF, Lovett GM, Fargione JE, Greenwood L, Serbesoff-King K, Randall JM & Fei S (2021) Insect and Disease Disturbances Correlate With Reduced Carbon Sequestration in Forests of the Contiguous United States. Front. For. Glob. Change 4:716582. [Volume 4 | Article 716582] doi: 10.3389/ffgc.2021.716582
Sitzia, T., T. Campagnaro, G. Brundu, M. Faccoli, A. Santini & B.L. Webber. 2021. Routledge Handbook of Biosecurity & invasive species. Chapter 7. Forest Ecosystems. ISBN 9780367763213
Weed, A.S., M.P. Ayers, J.A. Hicke. 2013. Consequences of CC for biotic disturbances in North American forests. Ecological Monographs, 83(4), 2013, pp. 441–470
Wu, H. 2023/24. Modelling Tree Mortality Caused by Ash Dieback in a Changing World: A Complexity-based Approach MSc/MPhil Dissertation Submitted August 12, 2024. School of Geography & the Enviro, Oxford University
Posted by Faith Campbell
We welcome comments that supplement or correct factual information, suggest new approaches, or promote thoughtful consideration. We post comments that disagree with us — but not those we judge to be not civil or inflammatory.
For a detailed discussion of the policies and practices that have allowed these pests to enter and spread – and that do not promote effective restoration strategies – review the Fading Forests report at http://treeimprovement.utk.edu/FadingForests.htm
an aye-aye – one of the highly endangered lemurs dependent on moist tropical forests of Madagascar; photo by Andrew Ciscel via Wikimedia
A forthcoming study examines two important issues: interactions of pathogens’ spread and changing climate, and invasive species threats to tropical islands’ forests.
Underwood et al. (in press) analyzed how an introduced vascular wilt pathogen — Leptographium calophylli – is likely to affect a tree endemic to Madagascar’s already threatened mid-level elevation humid & subhumid forests, Calophyllum paniculatum (sorry; I can find no photographs of the tree species).
Climate change is expected to cause substantial shifts in temperature and precipitation patterns on the island. These temperature and moisture regimes in turn govern pathogen sporulation, infection efficiency, and survival. They also affect the host’s levels of stress and defenses. The direction of change is not certain, however. In some cases, warming and other changes to the climate might facilitate a pathogen’s spread, allowing it to track shifts in the host’s range and expand into previously unoccupied refugia. In other cases, these changes might erect environmental thresholds that limit the pathogen’s survival and spread, thereby creating spatial refugia for the host.
diademed lemur, courtesy of Animalia
Environmental change increases the area of suitable landscape, that is, it weakens climatic barriers to establishment. Continued anthropogenic movement of some vector (biological or not) generates multiple introductory events over time. As a result, the likelihood of a successful establishment also increases, even if the probability per individual introduction is unchanged. Underwood et al. say that invasion outcomes thus become increasingly dependent on propagule pressure.
On many other tropical islands the threat from climate change is exacerbated by deforestation. On Madagascar, clearing driven by slash-and-burn agriculture and fuelwood harvesting has already reduced natural forest cover to less than 10% of its original extent. [For more on this topic, see e.g., Mittermeier et al. (2011).] Underwood et al. cite a determination by the ForestAtRisk model that humid forest in Madagascar could be almost entirely lost by 2100.
Loss of Madagascar’s forest has global implications. The island is one of 36 global biodiversity hotspots for both flora and fauna (e.g., lemurs). Its flora exceeds 12,000 plant species, of which 83% are endemic. In this case, the host tree species — Calophyllum paniculatum — is already considered vulnerable by the International Union for the Conservation of Nature (IUCN). Thus it is of global importance to understand the relative importance of several threats so that conservations can adopt the most effective countermeasures.
While they do not say so explicitly, it appears that Underwood et al. worry that too few of the conservationists active on Madagascar are paying attention to the possible impact of introduced pathogens. They note that pathogen-driven mortality of dominant or functionally unique trees can rapidly alter community structure and ecosystem function, potentially triggering local extinctions and cascading ecological consequences. For example, if an infection removes mature trees, their loss reduces fruit and nectar availability and so depresses populations of dependent wildlife. The trees’ death also diminishes above-ground carbon stocks and litter inputs. In combination, these impacts can shift community composition toward disturbance-tolerant states and heighten susceptibility at forest margins. These changes difficult to reverse once thresholds crossed.
red-bellied lemur in Ranomafana National Park – site of the first detection of Leptographium calphylli; via Flickr
This threat is not hypothetical. Since 2016 mature C. paniculatum at one site – a National Park – have been dying from a vascular wilt disease caused by a species in the Leptographium genus, probably Leptographium (formerly Verticillium) calophylli. While the species hasnot yet officially been recorded in Madagascar, it is established on neighboring Indian Ocean islands and across much of mainland Africa. Various species in the fungal genus are known to cause disease in other woody hosts. Underwood et al. suggest it was probably transported to Madagascar on infected wood, although they present no data.
Inside forests, Leptographium spp. are vectored by bark beetles in the Cryphalus genus. At least 25 Cryphalus species occur on the African Continent; some are vectoring disease on Seychelles and Mauritius.
The analysis by Underwood et al. indicates that future climatic conditions are likely to worsen the Leptographium calophylli infection over coming decades. The causal agent is likely to retain two-thirds of its current probable distribution and expand into previously uninhabited regions. The suitable habitat is expected to stretch across the entire north-south humid belt – the entire distribution of the host tree. Underwood et al. (in press) say it is even possible that the pathogen might remain in the forest, subsisting on other hosts, after C. paniculatum becomes functionally extinct across its range.
Meanwhile, that host – Calophyllum paniculatum – is projected to experience severe range shifts, with an overall net contraction across all climate change scenarios. It is forecast up to 67% of its current area by 2100. This range contraction will be compounded by fragmentation and dispersal limitation resulting from from deforestation. The refugia will be few and geographically isolated by late in the 21st century.
red-veined swallowtail; photographed in Ranomafana National Park by Frank Vassen, via Wikimedia
Are conservationists considering the implications of Leptographium calophylli’s probable persistence? Underwood et al. imply they are not; they say the impact of this and related pathogens on Madagascar & nearby islands is “still an unknown to the conservation community”. They urge their colleagues to conduct a set of research actions to identify, monitor, & limit the fungus’ spread – – and thereby improve the effectiveness of conservation efforts.
Host range & other targets: determine whether L. calophylli infects other taxa in Madagascar – especially the endemic species and genera. They suggest systematic field sampling of multiple species across sites within the core probable range of L. calophylli. A trained pathologists should be consulted to officially identify the pathogen.
Determine the spread phase of the pathogen. They suggest random sampling of species & sites within & outside of the fungus’ probable distribution, mapping the possible start point & dispersal patterns, including both anthropogenic & natural spread routes.
Assess applicability of IPBES tools & suggestions for invasive species management to the case of a fatal pathogen in the context of tropical islands’ characteristics. How might Madagascar implement prevention, early detection & rapid response systems?
I applaud Underwood et al. for trying to alert the conservation community active on tropical islands to the simultaneous impacts of multiple global & regional change drivers on vulnerable species. Probably other host-pathogen systems are experiencing the same diverging trajectories that might intensify their biodiversity loss, particularly when compounded by deforestation.
SOURCES
Mittermeier, R.A., E.E. Louis Jr., M. Richardson, C. Schwitzer, O. Langrand, A.B. Rylands. 2010. Lemurs of Madagascar. Conservation International, Arlington, USA. ISBN 9781934151235
Underwood, E.L., K.A Brown, A. Ronnfeldt, M. Mulligan, N. Walford, R. Allgayer. In press. Climate change facilitates fungal pathogen expansion while driving endemic host range contractions in a tropical biodiversity hotspot. Research Square.
Posted by Faith Campbell
We welcome comments that supplement or correct factual information, suggest new approaches, or promote thoughtful consideration. We post comments that disagree with us — but not those we judge to be not civil or inflammatory.
For a detailed discussion of the policies and practices that have allowed these pests to enter and spread – and that do not promote effective restoration strategies – review the Fading Forests report at http://treeimprovement.utk.edu/FadingForests.htm
The Trump Administration’s budget for Fiscal Year 2026 [which begins at the end of September 2025] proposes to eliminate funding for nearly all USFS research & Forest Health Protection.
Proposed Cuts to USFS Research: Timber the Sole Aim
In a letter from Office of Management and Budget (OMB) to Senate Appropriations Committee Chair Susan Collins (R-Maine, Director Russell Vought says the Administration wants to manage National forests “for their intended purpose of producing timber” and that the research and development program “is out of step with the practical needs of forest management for timber production.” The Administration proposes to eliminate funding for USFS research projects other than the small portion covering Forest Inventory and Analysis.
I understand that the USFS Chief told various NGOs that his job is to run the National Forest System, increase timber production by 40%, and do nothing else.
This single aim conflicts with the 1897 legislation founding and authorizing the USFS. It also violates provisions of subsequent legislation such as the Multiple-Use Sustained-Yield Act of 1960 and the National Forest Management Act of 1976. It also departs from long-standing US Forest Service policy – which is the intention.
The “intended purpose” of establishing “forest reserves” [which were later renamed National forests] has never been solely for timber production. The “Organic Act” of 1897 provided that any new forest reserves would have to meet the criteria of forest protection, watershed protection, and timber production.
Specifically, theORGANIC ACT OF 1897 [PUBLIC–No.2.] says:
“[All public lands heretofore designated and reserved by the President of the US under the provisions of the Act [of] March 3rd 1891, the orders for which shall be and remains in full force and effect, unsuspended and unrevoked, and all public lands that may hereafter be set aside as public forest reserves under said act, [these were the “forest reserves,”predecessors of “National Forests]” shall be as far as practicable controlled and administered in accordance with the following provisions:
“No public forest reservation shall be established, except to improve and protect the forest within the reservation, or for the purpose of securing favorable conditions of water flows, and to furnish a continuous supply of timber for the use and necessities of [US] citizens; but it is not the purpose or intent of these provisions, or of the Act providing for such reservations, to authorize the inclusion therein of lands more valuable for the mineral therein, or for agricultural purposes, than for forest purposes.”
The Department of the Interior, which then managed these forest reserves, promptly issued implementing regulations. The regulations stated that the “object” of forest reservations was:
“2. Public forest reservations are established to protect and improve the forests for the purpose of securing a permanent supply of timber for the people and insuring conditions favorable to continuous water flow.”
Therefore, I think the Administration has exaggerated the emphasis on timber production by calling it “the” intended purpose of the original establishment of National forests. The Administration has also chosen to ignore subsequent legislation, such as the Multiple-Use Sustained-Yield Act of 1960 and the National Forest Management Act of 1976.
Sec. 13 of the NFMA limits the sale of timber from each national forest to a quantity equal to or less than a quantity which can be removed from such forest annually in perpetuity on a sustained-yield basis. This limit might be exceeded under certain circumstances, but such excess must still be consistent with the multiple-use management objectives of the land management plan. Further, Sec. 14 requires public input into any decision to raise timber allowances.
During his period as Chief (1905 – 1910), Gifford Pinchot invented and applied the concept of “conservation” of natural resources. As a result “wise use” became accepted as the national goal.
Culminating more than a century of legislation and informed policy, the mission of the USDA Forest Service is to “sustain the health, diversity, and productivity of the nation’s forests and grasslands to meet the needs of present and future generations.”
Proposed Cuts to State, Private, and Tribal Forests
The budget also cuts $303 million from the State, Private, and Tribal Forests program. (I understand this zeroes out the entire program). The OMB Director alleges that the program has been “plagued by oversight issues, including allegation of impropriety by both the Agency and State governments.” I understand that this would eliminate the cooperative projects managed by the Forest Health Protection program, too.
Implications for Non-native Insects and Pathogens
Remember that USFS’s research and development program is intended to improve forest managers’ understanding of ecosystems, including human interactions and influences, thereby enabling improvements to the health and use of our Nation’s forests and grasslands. Most importantly to me, this program provides foundational knowledge needed to develop effective programs to prevent, suppress, mitigate, and eradicate the approximately 500 non-native insects and pathogens that are killing America’s trees.
The Forest Health Program provides technical and financial assistance to the states and other forest-management partners to carry out projects (designed based on the above research) intended to prevent, suppress, mitigate, and eradicate those non-native insects and pathogens. The program’s work on non-federal lands is crucial because introduced pests usually start their incursions near cities that receive imports (often transported in crates, pallets, or imported plants).
Eliminating either or both programs will allow these pests to cause even more damage to forest resources – including timber.
Both supporting research and on-the-ground management must address pest threats across all U.S. forests, including the more than 69% that are located on lands managed by others than the USFS. Already, the 15 most damaging of these pests threaten destruction of 41% of forest biomass in the “lower 48” states. This is a rate similar in magnitude to that attributed to fire (Fei et al. 2019). It is ironic that the Administration considers the fire threat to be so severe that it has proposed restructuring the government’s fire management structure.
I remind you that the existing USFS R&D budget allocates less than 1% of the total appropriation to studying a few of the dozens of highly damaging non-native pests. I have argued that this program should be expanded, not eliminated. Adequate funding might allow the USFS to design successful pest-management programs for additional pests (as suggested by Coleman et al.).
As a new international report (FAO 2025) notes, genetic resources underpin forests’ resilience, adaptability, and productivity. Funding shortfalls already undercut efforts to breed trees able to thrive despite introduced pests and climate change (the latter threat is still real, although the Administration disregards it). I have frequently urged the Congress to increase funding for USFS programs – which are sponsored primarily by the National Forest System and State, Private, and Tribal, although some are under the R&D program.
Please ask your Member of Congress and Senators to oppose these proposed cuts. Ask them to support continued funding for both USFS R&D and its State, Private, and Tribal Programs targetting non-native insects and pathogens. America’s forests provide resources to all Americans – well beyond only timber production and they deserve protection.
Contacting your Representative and Senators is particularly important if they serve on the Appropriations committees.
House Appropriations Committee members:
Republicans: AL: Robert Aderholt, Dale Strong; AR: Steve Womack; AZ: Juan Ciscomani; CA: Ken Calvert, David Valadao, Norma Torres; FL: Mario Diaz-Balart, John Rutherford, Scott Franklin; GA: Andrew Clyde; ID: Michael Simpson; IA: Ashley Hinson; KY: Harold Rogers; LA: Julia Letlow; MD: Andy Harris; MI: John Moolenaar; MO: Mark Alford; MS: Michael Guest; MT: Ryan Zinke; NC: Chuck Edwards; NV: Mark Amodei; NY: Nick LaLota; OH: David Joyce; OK: Tom Cole, Stephanie Bice; PA: Guy Reschenthaler TX: John Carter, Chuck Fleishmann, Tony Gonzales, Michael Cloud, Jake Ellzey; UT: Celeste Maloy; VA: Ben Cline; WA: Dan Newhouse; WV: Riley Moore
Democrats: CA: Pete Aguilar, Josh Harder, Mike Levin; CT: Rosa DeLauro; FL: Debbie Wasserman Schultz, Lois Frankel; GA: Sanford Bishop; HI: Ed Case IL: Mike Quigley, Lauren Underwood; IN: Frank Mrvan; MD: Steny Hoyer, Glenn Ivey; ME: Chellie Pingree; MN: Betty McCollum; NJ: Bonnie Watson Coleman NY: Grace Meng, Adriano Espaillat, Joseph Morelle; NV: Susie Lee; OH: Marcy Kaptur; PA: Madeleine Dean; SC: James Clyburn; TX: Henry Cuellar, Veronica Escobar; WA: Marie Gluesenkamp Perez; WI: Mark Pocan
Senate Appropriations Committee members:
Republicans: AK: Lisa Murkowski; AL: Katie Britt; AR: John Boozman (AR); KS: Jerry Moran; KY: Mitch McConnell; LA: John Kennedy; ME: Susan Collins; MS: Cindy Hyde-Smith; ND: John Hoeven; NE: Deb Fischer; OK: Markwayne Mullin; SC: Lindsey Graham; SD: Mike Rounds TN: Bill Hagerty; WV: Shelley Moore Capito;
Democrats: CT: Chris Murphy; DE: Chris Coons; GA: Jon Ossof; HI: Brian Schatz; IL: Richard Durbin; MD: Chris van Hollen; MI: Gary Peters; NH: Jeanne Shaheen; NM: Martin Heinrich; NY: Kirsten Gillibrand; OR: Jeff Merkley; RI: Jack Reed; WA: Patty Murray; WI: Tammy Baldwin
SOURCES
Coleman, T.W, A.D. Graves, B.W. Oblinger, R.W. Flowers, J.J. Jacobs, B.D. Moltzan, S.S. Stephens, R.J. Rabaglia. 2023. Evaluating a decade (2011–2020) of integrated forest pest management in the United States. Journal of Integrated Pest Management, (2023) 14(1): 23; 1–17
FAO. 2025. The Second Report on the State of the World’s Forest Genetic Resources. FAO Commission on Genetic Resources for Food and Agriculture Assessments, 2025. Rome.
Fei, S., R.S. Morin, C.M. Oswalt, and A.M. 2019. Biomass losses resulting from insect and disease invasions in United States forests. PNAS August 27, 2019. Vol. 116 No. 35 17371–17376
Posted by Faith Campbell
We welcome comments that supplement or correct factual information, suggest new approaches, or promote thoughtful consideration. We post comments that disagree with us — but not those we judge to be not civil or inflammatory.
For a detailed discussion of the policies and practices that have allowed these pests to enter and spread – and that do not promote effective restoration strategies – review the Fading Forests report at https://treeimprovement.tennessee.edu/
Guest blog by Kristy M. McAndrew, Department of Forestry, Mississippi State University
Virginia juniper (Juniperus virginiana) preforming its ecological role: succession in a field (in Ohio); photo by Greg Hume via Wikimedia
Spread of non-native species is a facet of global change that is an unintended consequence of the modern global trade network. Despite efforts put in place to limit such transport, such as International Standards for Phytosanitary Measures (ISPMs), unintentional spread of species continues, and thus, an important part of forest health research and management includes non-native monitoring and control efforts. As other aspects of global change, such as climate and weather patterns, shift, the dynamics between native landscapes and introduced pests may unexpectedly shift as well. For example, increased climate stress of tree hosts may weaken tree defenses, allowing species that historically have not been pests of concern to reach pest status.
Japanese cedar longhorned beetle (Callidiellum rufipenne; JCLB) is a wood boring beetle in the longhorned beetle family, Cerambycidae. The adults are reddish brown in color, and relatively small for longhorned beetles, at only around 1 cm in length. Japanese cedar longhorned beetle has a long history of establishing outside of its native range but has largely been considered a non-issue. It has long been disregarded as a pest because it feeds primarily on dead or dying trees in both the native and invaded ranges. However, there are more examples of these beetles feeding on stressed, but alive, trees in North America. Therefore, I think it is an important insect to take a closer look at.
Life cycle
These beetles have a one-year life cycle, most of which is spent inside a host tree. Adults emerge from host trees in the early spring and seek out other adults to mate with and trees to lay eggs on. Eggs are laid on thin parts of bark or in bark crevices, and when the eggs hatch larvae chew beneath the bark where they feed on the phloem until they have completed larval development. Once larvae are fully developed, they burrow further into the tree, into the xylem tissue, where they pupate, overwinter as fully formed adults, and continue the cycle the following spring.
Native range
The native range of JCLB is eastern Asia. It is common throughout the Korean peninsula and across the islands of Japan. It is also considered native to Eastern China and Russia. Within the native range JCLB is found primarily on dead and/or dying trees and is thus considered a secondary pest. On dead trees they can be found on any diameter of dead woody material, but on declining trees they will likely be in the small diameter branches and stems.
Arborvitae (Thuja occidentalis); photo by James St. John via Flickr
Invasion history
Japanese cedar longhorned beetle was first documented as an invasive pest in the early 1900s in France, and since then has established in at least fifteen countries (Clément 2023). Most of these countries are in Europe, but the United States and Argentina also have established populations. As with most woodboring insects, the invasion pathway is believed to have been wood packaging material being transported via global trade routes. Between 1914 and 2022 it was intercepted over 700 times (reviewed by KM). Since the implementation of ISPM No. 15, only six interceptions have been reported up to 2022 (USDA APHIS data reviewed by K.M.). [For Faith’s view on the regulation of wood packaging, see Fading Forests II and III (links provided at the end of this blog) and earlier blogs posted here under the category “wood packaging”. esp. 1 from 2015].
A USDA risk assessment completed in 2000 suggested other possible pathways of introduction, including balled nursery stock, green logs, and pruned branches (USDA APHIS and Forest Service, 2000).
In terms of establishments in North America, JCLB was first detected in natural forests in North Carolina in 1997. It was soon discovered in Connecticut in 1998; in neighboring New York in 1999; and in Massachusetts, New Jersey, and Rhode Island in 2000. It was quickly discovered feeding on live arborvitae (also called northern white cedar; Thuja occidentalis) in these invaded regions. JCLB has since been found in Pennsylvania (in 2010) and Maryland (in 2011). It is important to note that it is not clear when this species truly established, because of its previously discussed long history of being intercepted in ports of entry.
Most introduced populations of JCLB are found in either dead hosts or in the damaged/dead limbs of live hosts. In Buenos Aires, for example, storm-damaged trees with broken limbs are often where beetles are collected (Turienzo 2007). In the United States, eastern red cedar (Juniperus virginiana) and common juniper (Juniperus communis ) are the two native species most commonly affected, but so far there is no evidence of live trees of these species being infested (Maier 2007). However, a growing concern in the United States is that JCLB has been documented on live trees – particularly in urban environments. These trees are typically arborvitae, and they are typically stressed urban trees that have been overwatered and often show signs and symptoms of other health issues.
Host breadth
The host breadth of JCLB encompasses much of the family Cupressaceae. Maier (2007) identified 19 potential hosts from the literature and research, with the vast majority (14) of the hosts being Cupressaceae species, which is indicative of JCLB being a relative generalist, especially when considering species in the cypress family. This is important, because there are over 130 species within Cupressaceae worldwide that could be suitable hosts for JCLB, meaning host will not be a limiting factor in many invasion scenarios for this insect. Most often trees infested by JCLB need to be either stressed or dead, which limits suitability to an extent. However, many landscape trees are inherently stressed, whether it be from a history of roots being balled and wrapped in burlap, being planted in less than ideal scenarios, or being overwatered.
A few reports from research in Japan record JCLB feeding on plants in Pinaceae, primarily Pinus and Abies species. One article reports use of Larix kaempferi; another documented JCLB on the Taxaceae species, Taxus cuspidata. North American pine (Pinus spp.) and fir (Abies spp.) species have not been tested, but if they are revealed as suitable that would increase the availability of hosts in North America significantly.
In southern New England at least nine species have been confirmed as suitable, all of which are in the family Cupressaceae. Native and abundant junipers, such as Juniperus virginiana, appear to be highly suitable hosts. Additional host testing would be beneficial – especially Cupressaceae species that are either threatened or have a limited range. Within the United States there are a total of 28 native Cupressaceae species. Thus the suitable range (in terms of hosts) covers the entire Eastern half of North America through central Texas, most of the Pacific Coast, and widespread but spotty/disjunct areas throughout the Intermountain West and High Plains regions.
Atlantic white cedar swamp (Chamaecyparis thyoides) in Brendan Byrne State Forest, New Jersey; photo by Famartin via Wikimedia
Suitability
Tools such as environmental niche models can give helpful estimates of suitability. For species that are typically secondary pests, such as JCLB, it can be difficult to obtain non-biased data with good coverage to make reliable predictions. Preliminary research (unpublished) has been completed to estimate suitable habitat with limited occurrence records from the native range. Despite limited occurrences, models performed well and estimated moderate to high suitability in most temperate regions globally. These preliminary models are still being optimized by working with collaborators within the native range of JCLB to increase the number of occurrences. It is also important to note that these models are only accounting for climate data. Host data was not included, but Cupressaceae species are abundant globally, and therefore host availability is not likely a limiting factor for JCLB in establishing in regions.
Importance of monitoring species
While JCLB is still mostly limited to dead, dying trees, many of the species it may affect in the Eastern United States are already of heightened conservation concern. Wetland Cupressaceae, such as bald cypress (Taxodium distichum) and Atlantic White Cedar (Chamaecyparis thyoides), are valuable in terms of ecosystem services they provide in coastal, and inland, wetlands. These wetlands are encountering heightened stress in the form of increasing saltwater intrusion, increased storm strength, and changing landscapes, all of which may predispose trees to insect attack. Japanese cedar longhorned beetle has been successfully reared out of logs of Atlantic White Cedar, but thankfully has not been documented on live trees of this species (Maier 2009)[Ma1] . Bald cypress has not yet been tested for suitability. It is unknown if the stressors these trees are facing and will continue to face will impact JCLB’s ability to infest these landscapes, or if they will remain restricted to dead trees in these coastal forests. Regardless, given JCLB already has an established foothold in the Eastern United States, it is important to better understand the potential impacts of this insect.
First steps to understanding those impacts include 1) better documenting the host range in the regions and 2) determining the climate that may support the species. Hopefully we can continue research in these areas to best manage this non-native pest.
Much of the research conducted on JCLB in North America took place almost 20 years ago (Maier 2007, 2009), so updated sampling has potential to provide a wealth of information regarding spread rate, suitable climate, and establishment patterns.
bald cypress(Taxodium distichum); photo by Kej605 via Wikimedia; it is unknown whether this species is vulnerable to the Japanese cedar longhorned beetle
Sources
Clément F. 2023. Le point sur la distribution en France et en Europe de Callidiellum rufipenne (Motschulsky, 1861)(Coleoptera, Cerambycidae, Cerambycinae, Callidiini). Le Coléoptériste. 26(3):188–203.
Maier CT. 2007. Distribution and Hosts of Callidiellum rufipenne (Coleoptera: Cerambycidae), an Asian Cedar Borer Established in the Eastern United States. JOURNAL OF ECONOMIC ENTOMOLOGY. 100(4).
Maier CT. 2009. Distributional and host records of Cerambycidae (Coleoptera) associated with Cupressaceae in New England, New York, and New Jersey. Proceedings of the Entomological Society of Washington. 111(2):438–453. https://doi.org/10.4289/0013-8797-111.2.438
Turienzo P. 2007. New records and emergence period of Callidiellum rufipenne (Motschulsky, 1860) [Coleoptera:Cerambycidae: Cerambycinae: Callidiini] in Argentina. Boletín de Sanidad Vegetal, Plagas. 33:341–349.
United States Department of Agriculture Animal and Plant Health Inspection Service and Forest Service 2000. (Pasek, J.E., H.H. Burdsall, J.F. Cavey, A. Eglitis, R.A. Haack, D.A. Haugen, M.I. Haverty, C.S. Hodges, D.R. Kucera, J.D. Lattin, W.J. Mattson, D.J. Nowak, J.G. O’Brien, R.L. Orr, R.A. Sequeira, E.B. Smalley, B.M. Tkacz, W.W. Wallner) Pest Risk Assessment for Importation of Solid Wood Packing Materials into the United States. USDA APHIS and Forest Service. August 2000.
Posted by Faith Campbell
We welcome comments that supplement or correct factual information, suggest new approaches, or promote thoughtful consideration. We post comments that disagree with us — but not those we judge to be not civil or inflammatory.
For a detailed discussion of the policies and practices that have allowed these pests to enter and spread – and that do not promote effective restoration strategies – review the Fading Forests report at https://treeimprovement.tennessee.edu/
Scientists in New Zealand are saying explicitly that a forest’s unique mixture of species matters when considering the future. This mixture is the result of the forest’s evolutionary history. Losing members of the biological community reduces the forest’s ability to respond to current and future stresses – its resilience.
New Zealand’s forests are part of the broader legacy of the ancient supercontinent of Gondwanaland – the island nation’s plants have close relatives in South America, the Pacific Ocean islands, and Australia. Still, these forests are unique: 80% of New Zealand’s plant species are endemic. The forests are also species-rich. The warm temperate evergreen rain forests of the North Island are home to at least 66 woody plant species that can reach that reach heights above six meters (Simpkins et al. 2024).
These forests have been severely changed by human activity. In just ~ 750 years people have cut down approximately 80% of the original forest cover! (Simpkins et al. 2024) Of the eight million hectares of surviving native forest, a little over five million hectares is managed for the conservation of biodiversity, heritage, and recreation. Another 2 million hectares are plantations of non-native species.
sites in New Zealand where pine plantations are “wilding”
All these forests are challenged by introduced mammals – from European deer to Australian possums. Climate change is expected to cause further disturbance, both directly (through e.g., drought, extreme weather) and indirectly (e.g., by facilitating weed invasion and shifting fire regimes) (Simpkins et al. 2024).
Pathogen threats are also common threats to the native trees of the Pacific’s biologically unique island systems. For example, Ceratocystis lukuohia and C. huliohia (rapid ‘ōhi‘a death, or ROD). The latter is killing ‘ōhi‘a (Metrosideros polymorpha) on the Hawaiian Islands. More than 40% of native plant species in Western Australia are susceptible to Phytophthora cinnamomi. Here I focus on two pathogens, kauri dieback and myrtle rust, now ravaging New Zealand’s native flora. No landscape-level treatment is available for either pathogen.
When considering this suite of challenges, Simpkins et al. focus on these two pathogens’ probable impact on forest carbon sequestration. They worry in particular about erosion of the forests’ resilience due to loss of “ecological memory” – the life-history traits of the species (e.g., soil seed banks) and the structures left behind after individual disturbances.
one of the largest remaining kauri trees, “Tane Mahuta”, in Waipoua Kauri Forest; photo by F.T. Campbell
Kauri Dieback
The causal agent of Kauri dieback, Phytophthora agathidicida, is a soil-borne pathogen that spreads slowly in the absence of animal or human vectors. The disease affects a single species, Agathis australis (kauri, Araucariaceae). However, kauri is a long-lived, large tree that is a significant carbon sink. It probably modifies local soil conditions, nutrient and water cycles, and associated vegetation. Also, kauri has immense cultural significance.
Simpkins et al. note that kauri dieback threatens stand-level loss of A. australis – that is, local extinctions. In the absence of disturbance Kauri trees can grow to awe-inspiring size. In the 19th Century, before widespread logging, some were measured at 20 meters or more in circumference. Consequently, kauri dieback might cause a decline in aboveground live carbon storage of up to 55%. This loss would occur over a period of hundreds of years, not immediately.
Huge kauri are not likely to be replaced by other long-lived emergent conifers (based on an analysis of one species, Dacrydium cupressinum). Instead, kauri are probably going to be replaced by late-successional angiosperms. The authors discuss the ecological implications for levels of carbon storage and proportions of trees composed of Myrtaceae – exacerbating damage caused by myrtle rust (see below).
The expectation of Simpkins et al. that kauri will suffer at least local extinctions is based on an assumption that no kauri trees are resistant to the pathogen. Fortunately, this might not be true: different Agathis populations show various levels of tolerance to Agathis dieback. Identification and promotion of some levels of resistance could enable A. australis to retain a diminished presence in the landscape.
However, Lantham, et al. make clear that containing kauri dieback remains “challenging,” despite its discovery nearly 20 years ago (in 2006). Scientists and land managers have little information on the distribution of symptomatic trees, much less of the pathogen itself. This means they don’t know where infection foci are or how fast the disease is spreading.
As is often true, the pathogen is probably present in a stand for years, possibly a decade or more, before symptoms are noticed. This means that the current reliance on public reports of diseased trees, or targetting surveillance on easy-to-access sites (e.g., park entrances and along existing track networks), or at highly impacted areas readily identified through aerial methods, fails to detect early stages of infection. Indeed, it seems probable that P. agathidicida had been present in New Zealand’s ecosystems for decades before its formal identification.
The Waipoua forest is one of the largest areas of forest with old kauri stands in the country. A new analysis of aerial surveys done between 1950 and 2019, shows how the forest is changing. The number of dead trees increased more than four-fold and the number of unhealthy-looking trees increased 16-fold over these 70 years. Kauri dieback is now widespread in this forest, especially in areas near human activities like clearing for pasture or planting commercial pine plantations).
Lantham et al. have developed a model which they believe will help identify areas of higher risk so as to prioritize surveillance and inform responses. These could delimit the disease front and help implement quarantines or other measures aimed at limiting the spread of P. agathidicida to uninfected neighboring sites.
I hope New Zealand devotes sufficient resources to expand surveillance and management to levels commensurate with the threat to this ecologically and culturally important tree species.
Leptospermum scoparia; photo by Brian Gatwicke via Flickr
Myrtle Rust
Myrtle rust is a wind-borne disease that affecting numerous species in the Myrtaceae, including some of the dominant early successional species (e.g., Leptospermum spp.). Simpkins et al. expect that myrtle rust might hasten the decline of two such tree species (L. scoparium and Kunzea ericoides). However, these trees’ small size and rapid replacement by other species during succession minimizes the effect of their demise on carbon storage.
Because I am concerned about the irreplaceable loss to biodiversity, I note that Simpkins et al. also feared immediate threats to some trees in the host Myrtaceae family, specifically highly susceptible species such as Leptospermumbullata.
As I reported in a recent blog, a second group of scientists (McCarthy et al.) explored the threat from myrtle rust more broadly. Austropuccinia psidii has spread through Myrtaceae-dominated forests of the Pacific islands for about 20 years.
Trees in the vulnerable plant family, Myrtaceae, are second in importance (based on density and cover) in New Zealand’s forests. Successional shrub communities dominated by the two species named above, Kunzea ericoides and Leptospermum scoparium, are widespread in the northern and central regions of the North Island and in northeastern and interior parts of the South Island. These regions’ vulnerability is exacerbated by the area’s climate, which is highly suitable for A. psidii infection (Simpkins et al. 2024).
McCarthy et al. concluded that ifLeptospermumscoparium and Kunzea ericoides prove to be vulnerable to myrtle rust, their loss would cause considerable change in stand-level functional composition across these large areas. They probably would be replaced by non-native shrubs, which are already common on the islands. Any resulting forest will differ from that formed via Leptospermeae succession.
These authors also worry that the risk to native ecosystems would increase if more virulent strains of the myrtle rust pathogen were introduced or evolved. They note that A. psidii is known to have many strains and that these strains attack different host species.
SOURCES
Latham, M.C., A. Lustig, N.M. Williams, A. McDonald, T. Patuawa, J. Chetham, S. Johnson, A. Carrington, W. Wood, and D.P. Anderson. 2025. Design of risk-based surveillance to demonstrate absence of Phytophthora agathidicida in New Zealand kauri forests. Biol. Invasions (2025) 27, no. 26
McCarthy, J.K., S.J. Richardson, I. Jo, S.K. Wiser, T.A. Easdale, J.D. Shepherd, P.J. Bellingham. 2024. A Functional Assessment of Community Vulnerability to the Loss of Myrtaceae from Myrtle Rust. Diversity and Distributions, https://doi.org/10.1111/ddi.13928
Simpkins, C.E., P.J. Bellingham, K. Reihana, J.M.R. Brock, G.L.W. Perry. 2024. Evaluating the effects of two newly emerging plant pathogens on North Aotearoa-New Zealand forests using an individual-based model. Ecological Modelling, www.elsevier.com/locate/ecolmodel
Posted by Faith Campbell
We welcome comments that supplement or correct factual information, suggest new approaches, or promote thoughtful consideration. We post comments that disagree with us — but not those we judge to be not civil or inflammatory.
For a detailed discussion of the policies and practices that have allowed these pests to enter and spread – and that do not promote effective restoration strategies – review the Fading Forests report at https://treeimprovement.tennessee.edu/
Europe has been invaded by two insect species that North Americans should be watching out for. First, a Cerambycid, the wasp-mimicking tiger longicorn beetle, Xylotrechus chinensis. And second,the Buprestid cypress jewel beetle, Lamprodila festiva. We should also ensure that none of the other 500+ beetles introduced to Europe poses a threat to our trees. These are summarized in a 2024 paper by Bunescu et al.
Tiger Longicorn Beetle
This beetle is native to eastern Asia. It feeds on and kills mulberry trees (Moraceae: Morus spp.). It might also attack apple and pear trees and grapevines – Asian sources report these as hosts. The status of grapevines has been questioned by a Spanish experiment, in which artificial inoculations failed. I have seen no further information about the vulnerability of apple (Malus spp.) and pear (Pyrus spp.) (Saarto i Monteyu, Costa Ribeu, and Savin 2021)
In Europe, the pest threatens mulberry trees which are commonly planted for shade and ornamentation, especially in southern France, Spain and Greece (Saarto i Monteyu, Costa Ribeu, and Savin 2021). For example, there are more than 20,000 mulberry trees in Athens (EFSA 2021). The trees’ abundance contributes to spread of any associated pests, the level of damage caused by falling branches, and the expense of tree removal. Economic damages are those typically associated with wood-borer invasions of urban areas. That is, the cost of tree removals, loss of shade and amenity values, and increased risk of injury from falling branches.
We Americans should be concerned, too. Wild red mulberry (Morus rubra) occupies much of the eastern United States, from southern New England west to southeastern Minnesota, then south along the eastern edge of the Great Plains to central Texas, and east to southern Florida. It is also found in Bermuda. It grows primarily in flood plains and low moist hillsides. . Presumably it would also be attacked by Xylotrechus chinensis, although I don’t know whether anyone has tested this. As a native tree, red mulberry plays a role in natural ecosystems, including wildlife food supplies. Thus, America would see even more significant losses if Xylotrechus chinensis were to establish.
Morus rubra in Fairfax County, Virginia; photo by Fmartin via Wikipedia
Red mulberry is already declining in parts of its central range, possibly due to a bacterial disease. The effects and extent of this disease have not been investigated thoroughly.
Apples and pears are important crops across North America; the farm-gate value is estimated at $3.2 billon.
Introductions of the beetle to Spain, France, and Greece might have resulted from inadequately-treated wood packaging or other wood products. Detections of the species in wood imports were reported in Germany in 2007 and 2017 (Saarto i Monteyu, Costa Ribeu, and Savin 2021). The U.S. has also intercepted X. chinensis at least once, at the port of Philadelphia, in 2011 (EFSA 2021).
These detections have raised questions to which no-one yet has answers. First, can X. chinensis develop in cut logs? The European Food Safety Agency concluded that it can (EFSA 2021). Second, one detection involved a shipment of wooden items made from birch (Betula spp.) and willow (Salix spp). It is not yet clear whether these taxa are also hosts (EFSA 2021). (The wood species were not specified in the case of the other interceptions.) I have blogged often about how “leaky” the wood packaging pathway has been; to see these blogs, scroll below the “archives” section of the webpage, then click on the category “wood packaging”.
European scientists believe X. chinensis might also be transported in shipments of plants for planting. However, the beetle prefers to oviposit on large trees. This pathway is less viable for the United States since USDA APHIS allows imports of mulberries (Morus) and pears (Pyrus) only from Canada. Apple trees (Malus spp.), however, may be imported from France – which hosts an introduced population of X. chinensis – and other European countries.
Detection of any invasion by X. chinensis will pose the usual difficulties associated with woodborers. In some European cities, hundreds or even a thousand trees were infested before the outbreak was detected (EFSA 2021).
I am concerned that the Europeans appear to have been slow to respond to the threat from Xylotrechus chinensis. After several outbreaks were discovered in Greece, France, and Spain in 2017 and 2018, the European and Mediterranean Plant Protection Organization (EPPO) added X. chinensis to its Alert List. This action requires member states (which are not limited to European Union members) to report new outbreaks and inform about efforts to either stop or eradicate them (Saarto i Monteyu, Costa Ribeu, and Savin 2021).
Shortly afterwards the European Union Commission requested the European Food Safety Agency (EFSA) to conduct a risk assessment. This analysis was completed in 2021. (It contains lots of photos of the insect and its damage.) The analysis concluded that Xylotrechus chinensis could probably infest most areas in the Union and cause significant damage. The species meets the criteria for designation as a quarantine pest in the Union. However, as of December 2024, this action had not been taken. As a result, control measures for this species are not mandatory.
Introductions continue; an outbreak in Lombardy, Italy, was found in June 2023 (Sarto i Monteys, Savin, Torras i Tutusaus & Bedós i Balsach 2024). European regulations – following IPPC standards – also are linked to named pests and known outbreak locations. Such restrictions almost guarantee that the pest will continue to spread from not-yet-detected outbreaks. (Decades ago, after the emerald ash borer invasion, Michigan’s State Plant Regulatory Official, Ken Rasher, noted that, to be effective, “slow the spread” efforts must apply to areas beyond the known limits of the pest’s range.) The EFSA risk assessment did suggest delimitation of buffer zones around known European outbreaks. I don’t know whether such zones have been set up.
The risk assessment also recommended [true?] improving detection of this insect by developing male pheromones as lures. These have not been acted on. Guidance on best timing for treatment [trunk injections of systemic insecticides] also appears to have been taken up by Greece but not by Spain (Sarto i Monteys, Savin, Torras i Tutusaus & Bedós i Balsach 2024).
These authors include more information about the Xylotrechus chinensis life cycle and trajectory of the invasion,. They note that climate change appears to be altering the insect’s phenology. Especially, the adult flight period is beginning earlier in the spring.
Lamprodila festiva; Udo Schmidt via flickr
Cypress jewel beetle
This second pest of concern is a buprestid that attacks trees in the Cupressaceae. Infested trees generally die within a few years.
In its native Mediterranean range, the beetle feeds on native Juniperus, Cupressus and Tetraclinis. In invaded urban landscapes of Europe it attacks primarily introduced Cupressaceae , particularly Thuja, Chamaecyparis, Platycladus, Callitris, and some hybrids (Cupressocyparis). It has also been recorded as damaging Sequoia sempervirens (Brunescu, et al., 2024). (Genera in bold are native to North America.)
Thuja occidentalis; photo by H. Zell via Wikimedia
White cedar, Thuja occidentalis is the focus of Brunescu, et al.’s article. It is native to eastern Canada and much of the north-central and northeastern United States. The European and Mediterranean Plant Protection Organization (EPPO) has identified eight species in the Lamprodila genus as important pests, (Brunescu et al. 2024) so the danger might be more widespread. The invasion of Europe is probably the result of adult flight or other short-range transport. The article does not suggest pathways that the species might exploit to cross oceans.
SOURCES
Bunescu, H., T. Florian, D. Dragan, A. Mara, I-B. Hulujan, X-D. Rau. 2024 The Cypress Jewel Beetle Lamprodila Festiva Linné, 1767 (Coleoptera: Buprestidae), an Invasive Major Pest of Thuja Occidentalis Linné in Romania Hop and Medicinal Plants, 2024 XXXII, No. 1-2, 2024.
Saarto i Monteyu V., A. Costa Ribeu. I. Savin. 2021a. The invasive longhorn beetle Xylotrechus chinensis, pest of mulberries, in Euro: Study on its local spread & efficacy of abamectin control Plos One January 29, 2021. https://doi.org/10.1371/journal.pone.0245527
Sarto i Monteys, V., I. Savin, G. Torras i Tutusaus & M. Bedós i Balsach. 2024b. New evidence on the spread in Catalonia of the invasive longhorn beetle, Xylotrechus chinensis, & the efficacy of abamectin control. Scientific Reports | (2024) 14:26754 | https://doi.org/10.1038/s41598-024-78265-xwww.nature.com/scientificreports/
Posted by Faith Campbell
We welcome comments that supplement or correct factual information, suggest new approaches, or promote thoughtful consideration. We post comments that disagree with us — but not those we judge to be not civil or inflammatory.
For a detailed discussion of the policies and practices that have allowed these pests to enter and spread – and that do not promote effective restoration strategies – review the Fading Forests report at https://treeimprovement.tennessee.edu/
