Southern Hemisphere forests: unique biomes face high invasion risk

native Olia woodiana in forest of Kwa-Zulu-Natal, South Africa. Photo by MJK via Wikimedia

The forests of the Southern Hemisphere faces an elevated risk from introductions of tree-killing arthropods and pathogens. I have already blogged about individual threats to the trees of South Africa, Australia, South America, and Madagascar. This blog will add information about those invasions and provide an overview and call for action.

Burgess and Winfield (2026) [full citation at the end of this blog] – eminent pathologists from South Africa – warn their colleagues that the economic and biological barriers that have long protected the biomes of the Southern Hemisphere from bioinvasion have now collapsed. While these countries remain geographically distant from each other and the Northern Hemisphere, and their floras are certainly still unique, they are now participating in global trade. In addition, they increasingly rely on industrial plantation forestry dominated by non-native tree species – some of them from the North. Finally, climate change is putting additional stress on all these systems.

Burgess and Winfield (2026) advise their colleagues to learn lessons from the alarming history of bioinvasion in the Northern Hemisphere. Over the past century, these forests have been battered by a series of invasions by tree-killing pathogens. They mention the examples we know too well: chestnut blight (Cryphonectria parasitica), white pine blister rust (Cronartium ribicola), “Dutch” elm disease (Ophiostoma ulmi/Ophiostoma novo-ulmi), pine wilt nematode (Bursaphelenchus xylophilus), sudden oak death (Phytophthora ramorum), ash decline (Hymenoscyphus fraxineus), laurel wilt (Harringtonia lauricolia), ʻōhiʻa or myrtle rust (Austropuccinia psidii), https://www.dontmovefirewood.org/pest_pathogen/ohia-rust-html/ and rapid ʻōhiʻa death (Ceratocycstis lukobia and huliohula). They do not address the similarly depressing history of arthropod interchanges among these realms.

Burgess and Winfield (2026) warn that these invasions are not anomalies but rather predictable outcomes of global commerce – especially in “plants for planting” – interacting with the ecological vulnerability arising from the biogeographic similarity of these temperate and boreal forests. The Northern Hemisphere is dominated by a single flora (the Holarctic).

The Southern Hemisphere has an astounding seven floral kingdoms, distributed as follows: Chile-Patagonian, Neotropical, African, Indo-Malesian, Australian, Novozealandic, and Cape. These regions were formerly part of Gondwanaland. Each has an exceptionally high level of endemism (Burgess and Wingfield 2026), but with phylogenic relationships to the flora of other regions. Changing trade patterns are opening introductory pathways for exchange of tree pests hosted by related but geographically distant species. The regional floras’ uniqueness multiplies the conservation importance of protecting them. I summarize Southern Hemisphere countries’ levels of endemism at the end of this blog.

The lessons Burgess and Wingfield (2026) draw from the bioinvasion history of the Northern Hemsphere include:

1) Trade in living plants – and even seeds – creates a high risk that accompanying microbes will be introduced. As noted by others – e.g., Tanney et al. (2025) – members of each plant species’ “phytobiome” are ubiquitous, diverse, difficult to detect, and sometimes capable of causing serious diseases, especially in naïve hosts.

2) After introduction, these microbes can evolve rapidly to adapt to new environments or hosts. They often become more aggressive and difficult to manage in the process. Plant health officials need to anticipate this evolution.

3) Impacts can be catastrophic when a dominant tree species is killed. When chestnut blight killed the vast majority of chestnut trees in North America it destabilized ecosystems and reshaped successional trajectories.

4) The threat is magnified when the receiving ecosystem is a forestry plantation because of the extensive monocultures.   

Building on Burgess and Wingfield (2026) and other sources (where noted), I provide the following summary of threats to native and planted forests in these regions. Some, e.g., Phytophthora cinnamomi and Sirex noctilio, were introduced many decades ago. But too many are recent arrivals.

I. Threats to tree species native to the region.

1) Australia

trees killed by Phytophthora cinnamomi; photo by Western Australia Parks & Wildlife
  • Carnegie et al. (2026) report that more 300 non-native insect pests, pathogens and nematodes are established on tree or shrub hosts in Australia. Twenty percent have caused moderate to high impacts to commercial plantations, urban forests, or trees in natural ecosystems.
  • Phytophthora cinnamomi: As noted, this pathogen was introduced decades ago. It has devastated jarrah (Eucalyptus marginata) forests and diverse plant communities. It causes severe mortality in Proteaceae, Epacridaceae, and Fabaceae. Of ~5700 described plant species in the state of Western Australia, 40% ( >2,300) can be killed (Carnegie et al. 2026).
  • Austropuccinia psidii (myrtle rust): threatens trees in the Myrtaceae worldwide. The rust has spread widely recently, causing severe ecological impacts in invaded regions. The “epidemic strain” introduced to Australia (and other regions) does not infect the commercially important genus Eucalyptus, which probably explains Australian authorities’ lackluster response. However, other lineages of the pathogen found in South America do infect Eucalyptus; introduction of one of these to Australia could be disastrous. Cautionary example of how rapidly a plant pathogen can conquer new ecosystems (Burgess and Wingfield 2026). Already, plant 76 species have been determined to be at risk to myrtle rust. The early focus has been on understory species in the eastern rainforests. Now that myrtle rust has been detected in Western Australia, a biodiversity hotspot with more than a thousand Myrtaceae taxa, most of which are expected to be susceptible. [Summary of Proceedings: Australian Myrtle Rust Conference Sydney, June 2023]
  • Polyphagous shot hole borer Euwallacea fornicatus s.s. (PSHB) and associated pathogen F. euwallacea were detected in 2021, probably three years after it established in Western Australia. Nearly 9% of hosts are regionally indigenous species, 21.3% native to other parts of the continent. There will probably be a considerable impact on the health of conservation reserves & national parks in the future. After trying for four years to eradicate the population, in November 2025, the National Management Group began a transition to a Management Response Plan. This new approach will impose significant additional expenses on local governments and communities for tree removal and replacement (Dell, Xu, & Chi 2026). Meanwhile, the state of New South Wales (2,000 miles away on the other side of the continent) initiated surveillance for PSHB. An assessment determined that 47% of urban trees in Sydney are susceptible to PSHB (Carnegie et al. 2026).
  • Erythina gall wasp Quadrastichus erythrinae (Carnegie and Nahrung 2019). Although Australia is home to at least one native species in the Erythrina genus, E. vespertilio, this pest has not been included on the environmental pest watch list. 

2) New Zealand

native forest on North Island, New Zealand, with kauri tree; photo by F.T. Campbell
  • Kauri dieback / Phytophthora agathidicida is killing Agathis australis, a keystone species that shapes forest structure and has enormous cultural importance for the Maori.

3) South Africa

  • Phytophthora cinnamomi — in the country since 1931 — threatens ecosystem collapse in Cape Floral Kingdom (Burgess and Wingfield 2026).
  • Polyphagous shot hole borer Euwallacea fornicatus s.s. and associated pathogen F. euwallacea The largest outbreak of Fusasrium disease in the world: it is present in every province except Limpopo. The Disease is recorded on 162 tree species. Seventy-eight of these species are indigenous to the country. Eighty-four are “competent” hosts and 78 are “Fusarium colonised” hosts (Townsend, Hill, Hurley and Roets 2025).
  • Fungus Seiridium neocupressi detected on the native tree, Widdringtonia nodiflora year The two other species in the genus, W. wallichii and W. schwartzii, occur in small endemic and threatened populations (Wingfield et al. 2022).

4) Madagascar

  • Leptographium calophylli killing an endemic tree species in mid-level elevation humid and subhumid forests, Calophyllum paniculatum.  

5) South America

  • Polyphagous shot hole borer Euwallacea fornicatus s.s. and associated pathogen F. euwallacea has been detected in Brazil, Argentina, and Uruguay. Dell, Xu, & Chi (2026) report that in Uruguay a significant number of native species are reproductive hosts for the beetle. Models developed by Coates and Philips (2026) indicate PSHB could thrive in more suitable habitats.
  • One of the continent’s native conifers, Austrocedrus chilensis is under attack by both an introduced aphid, Cinara cupressi and the pathogen Phytophthora austrocedri
Chilean cypress; photo by LBM via Wikipedia

Lantschner et al. (2026) identified seven independent introduction events of E. fornicatus: to the continental U.S., Hawai`i, South America, Central Europe, Spain, Türkiye, & South Africa.

II. Threats to plantations:

1) Co-evolved insects and pathogens have followed their Pinus hosts to plantations in the Southern Hemisphere due to biosecurity failures

  • Insects: Sirex noctilio (decades ago); recently Sirex obesus in Brazil (2023); Orthotomicus erosus and Cyrtogenius luteus (Stazione, Soliani, and Cognato 2026).
  • Pathogens: Fusarium circinatum, Dothistroma septosporum, Diplodia sapinea,  Phytophthora pinifolia, Lecanosticta acicula.

2) Because the Myrtaceae family is distributed around the Southern Hemisphere, plantations of Eucalyptus trees are now being damaged by both pathogens from their native Australian range (Teratosphaeria spp.) and local pathogens that have switched hosts (Chrysoporthe cubensis and Chrysoporthe deuterocubensis).

Burgess and Wingfield (2026) express fear that another set of pathogens, rapid ohia death (Ceratocystis lukuohia and Ceratocystis huliohia) might spread from Hawai`i farther across Oceania to New Zealand and other locations with important species in the Metrosideros genus. I note that the more virulent (on ʻōhiʻa) C. lukuohia belongs to an Asian-Australian clade (Luiz et al. 2023).  

Australia

I blogged in November 2022 about how surprising tardy Australia was in implementing phytosanitary programs to protect the nation’s forests. Scientists had spent 30 years trying to get such a program implemented. Burgess and Wingfield (2026) point out that 16 of 17 pathogens detected in Australia over the period 1996 – 2017 had established; several had had major impacts. None has been eradicated. They also note that the introduction of Austropuccinia psidii exposed significant weaknesses in post-border surveillance and response. 

The Australian federal govt has implemented a detection trapping program targeting the Asian spongy moth (= flighted spongy moth complex) at the 4 major ports (i.e., Melbourne, Hastings, Geelong and Portland) since 1996.

Carnegie et al. (2026) describe some helpful steps. They focus on establishment in 2022 of a National Forest Pest Surveillance Program (now called Forest Watch Australia). The program is funded by a levy on forest plantation growers and State governments, based on the relative risk to each region. Surveillance efforts are concentrated in areas deemed high risk areas for entry and establishment of forest pests. Program is coordinated through the federal agency, Plant Health Australia. It provides professional training for surveillance and diagnostic staff on an annual basis. Staff also conduct risk modelling and analysis to identify high-risk areas for surveillance in all major capital cities — Sydney, Melbourne, Brisbane, Hobart, Adelaide, Perth and Darwin.

The Australian program encounters the universal challenges regarding sustained resources, the diagnostic burden placed on scientists, and ownership of data. Carnegie et al. (2026) and Burgess and Wingfield (2026) discuss the growing availability of new tools for detection, e.g., molecular diagnostic tools.

Response options are guided by the national Emergency Plant Pest Response Deed and PlantPlan,  which outlines agreed government and industry responsibilities and cost-sharing protocols.  

Carnegie et al. (2026) praise especially the sentinel tree component of the surveillance program. Participants include governments at the federal, state, and municipal levels; botanical gardens; and industry. In 2024/25, trained staff inspected 5,136 trees of more than 50 species across 45 genera nationwide. Thanks to this expanded effort, the detection rate of new species has doubled since 2018, from ~ 1.5 to ~ 3 / year. Still, in three of four cases discussed by Carnegie et al. (2026), the introduced organisms were determined to be too widespread for eradication to be successful. They cite the PSHB outbreak in Perth as an example. Carnegie et al. (2026) advocate strengthening sentinel tree surveillance in urban landscapes so detection will occur sufficiently early to allow consideration of initiating an eradication or containment response. One worrisome aspect: there is no requirement to conduct more thorough surveillance or diagnostics on a species detected in the country once it has been ruled not “of concern”.

Burgess and Wingfield (2026) advocate embracing broader preventative approaches that address the underlying mechanisms of pathogen emergence and spread. I forthcoming blog will discuss proposed new approaches.

Background: Southern Hemisphere Flora

South Africa

As I blogged recently, South Africa’s flora is diverse (more than 20,000 indigenous species) and has high levels of endemism. That blog did not discuss the Myrtaceae family – my focus now because of the spread of the pathogen Austropuccinia psidii. According to Braam van Wyk and Hugh Glen (pers. comm.), the country has 24 native Myrtaceous species. These include 14 species in the Eugenia genus, six in the Syzygium genus. The one Metrosideros grows in the Cape Floristic Region. All the others occur in forest or grassland along the Indian Ocean Coastline well into Mozambique and further inland – some into Botswana and Zimbabwe.

Australia                      

Australia ranks second on Earth in the proportion of its flora that is endemic: 88% of plants. Australia comprises the native range of 38% of all Myrtaceae species on Earth, and of 66% of genera in the family (Brett Summerell, 2023 workshop). These 87 or 88 genera contain ~ 2,250 species and subspecies (Makinson 2018). This plant family constitutes ~10% of the continent’s native flora. They occur in 11 of 13 major vegetation formations (Carnegie et al.2026). The southwestern corner of Western Australia alone has ~5700 described plant species (Carnegie et al 2026).

Rhodomyrtus psidioides – one of the Australian Myrtaceae under greatest threat from myrtle rust; photo by Zaadero via Wikimedia

The island of Madagascar is home to 12,000 plant species, of which 83% are endemic.

SOURCES

Burgess T.K., and M.J. Wingfield. 2026. Unveiling a Hidden Menace: Invasive Tree Pathogens, Less Known but Increasingly Threatening Southern Hemisphere Forests Annual Review of Phytopathology #s

Carnegie AJ, Summerell BA, Trollip C, Tovar F, Smith DI and McDonald J (2026) Sentinel trees for early detection of non-native forest pests and pathogens in AU. Front. For. Glob. Change 9:1801183. doi: 10.3389/ffgc.2026.1801183

Ceriani-Nakamurakare, E., Johnson, A.J. and Gomez, D.F. (2023) Uncharted Territories: First report of Euwallacea fornicatus (Eichhoff) in South America with new reproductive hosts records. Zootaxa, 5325 (2), 289–297. https://doi.org/10.11646/zootaxa.5325.2.10

Lantschner, M.V., Ceriani-Nakamurakare, E., Johnson, A.J. et al. Invasion history reconstruction and potential distribution of the ambrosia beetles Euwallacea fornicatus and E. perbrevis, two global emerging pests. J Pest Sci 99, 100 (2026). https://doi.org/10.1007/s10340-026-02070-w

Luiz, B.C., C.P. Giardina, L.M. Keith, D.F. Jacobs, R.A. Sniezko, M.A. Hughes, J.B. Friday, P. Cannon, R. Hauff, K. Francisco, M.M. Chau, N. Dudley, A. Yeh, G. Asner, R.E. Martin, R. Perroy, B.J. Tucker, A. Evangelista, V. Fernandez, C. Martins-Keli’iho.omalu, K. Santos, R. Ohara. 2023. A framework for establishlishing a rapid ‘Ohi‘a death resistance program  New Forests 54, 637–660. https://doi.org/10.1007/s11056-021-09896-5

Stazione, L., Soliani, C., Cognato, A. et al. Reconstructing the invasion history of the bark beetles Orthotomicus erosus and Cyrtogenius luteus (Coleoptera, Curculionidae, Scolytinae) in South America. Biol Invasions 28, 49 (2026). https://doi.org/10.1007/s10530-026-03779-6

Tanney, J.B., M. Kemler, M. Vivas, M.J. Wingfield, and B. Slippers. 2025. Silent invaders: the hidden threat of asymptomatic phytobiomes to forest security. New Phytologist (2025) 247: 533–545 doi: 10.1111/nph.70209

Wingfield, M.J. S. Marincowitz, N.Q. Pham, F. Roets, T. Paap, B.D. Wingfield, J. Aylward. 2022. Cypress canker: An important disease discovered for the first time on a native South African tree. Plant Pathology 2022;71:1735-1742

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

Or

https://fadingforests.org

Tree regeneration update

sycamores overcome by invasive vines, primarily porecelain berry; photo by F.T. Campbell

18 months ago I blogged about the impacts of invasive plant species and deer on regeneration of tree species.  I provide more information on some National Park units’ efforts to counter these threats – contemporaneous with release of the report.

Emma Brentjens has posted a report with pictures. She focuses on National parks in the National Capital Region (NCR). These are the National parks in Maryland, Virginia, and a small part of West Virginia, including Catoctin Mountain Park; Antietam, Manassas, and Monocacy National battlefields; Harpers Ferry and Chesapeake & Ohio National Historical parks; Rock Creek Park; some parkways; and some urban parks, e.g. Anacostia and East Potomac. Forests constitute three-quarters of the landcover in National parks of this region. Regional NPS staff noted poor regeneration as early as 2006. In the years since, most have ramped up IAS plant management, & begun deer population management. As result, the previously steady regeneration declines have begun to reverse.

They analyze forest vegetation data using the Stocking Index, which calculates seedling density per hectare. “Acceptable” forest regeneration is defined as 67% of plots being adequately stocked. Unfortunately – from my perspective – the program accepts non-native tree species as successful regeneration.

Comparing the parks’ Stocking Index in 2006-2009 to 2020-2024 showed that most of these parks have seen an increase in the Stocking Index, including most sites managing deer populations. Still, no parks have yet achieved adequate stocking for more than 30% of the plot.

Most parks that are managing their deer populations show growth in seedling density, e.g., Catoctin Mountain Park (illustrated in the report); Harpers Ferry National Historical Park; and Monocacy National Battlefield. In Rock Creek Park, which also manages deer, many seedlings have grown into saplings.

deer in Shenandoah National Park; photo by F.T. Campbell

Most of the sites with declines in seedling density have not implemented deer management. These include several segments of Chesapeake & Ohio Canal National Historical Park and the parkways. Unfortunately, it also includes Prince William Forest Park, which at 15,000 acres is the largest protected area near Washington, D.C.

Brentjens notes that while these early results are promising, seedling levels are highly variable and do not guarantee forest regeneration. Continued management is necessary to sustain improvements.

As part of deer management, the NP units are also monitoring spread of chronic wasting disease (CWD) among white-tailed deer. Current CWD rates are very low (< 1%). The disease has been confirmed at four parks in the region: Antietam, Manassas, and Monocacy National battlefields and Harpers Ferry National Historical Park.

(Chronic wasting disease is not an effect too to reduce deer populations. First, is relying on such suffering an ethical approach? Second, significant declines in deer populations are not observed until infection rates reach about 29% of the population. Parks in the National Capital Region are not expected to reach that level for at least 20 years.

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

Or

https://fadingforests.org

Protecting forests’ ecosystem values: minimize disturbance & curtail nearby planting of non-native species

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

Or

https://fadingforests.org

Science advances re: ALB & EAB – cause for hope?

Science continues to provide new hope for management of some of the tree-killing insects introduced to North America.

searching for ALB; photo by USDA

Asian longhorned beetle

USDA Forest Service and APHIS scientists have developed an Asian Longhorned Beetle (ALB; Anoplophora glabripennis) Eradication and Risk Tracking (ALBERT) 1.0 toolkit. Trotter et al. (2024) believe that the toolkit will improve the efficiency of ALB eradication programs by better estimating changing infestation risk in real time and comparing efficacy of various management strategies in those circumstances. Managers can thus set priorities for surveys and eradication efforts, optimize surveys, and improve eradication outcomes. Improved efficiency will also reduce long-term management costs.

Data on dispersal and infestation risk are analyzed at the hectare level. Inputs are contemporary, so they reflect hectare-scale changes in risk level and ALB detectability simultaneously with program implementation. Managers and stakeholders can track and visualize the distribution of risk on the landscape. In the example, the risk of infestation is shown by the color and height of the surface on the landscape.

I note that the research was completed two years ago. I wish the USFS “Rooted in Research” brief had discussed whether programs managing extant ALB infestations are using the tool – and finding it useful.

Suppressing emerald ash borer populations

black ash swamp via Flickr

Minnesota authorities and scientists wish to protect the expansive black ash (Fraxinus nigra) swamps located mostly in the central & northern portions of the state. There, the emerald ash borer (EAB; Agrilus planipennis) has not yet spread to all vulnerable locations. They consider that use of pesticides is not concerned feasible in these locations. And two of the introduced biocontrol agents have limited efficacy. One of the wasps—Tetrastichus planipennisi— can oviposit in EAB larvae only in trees with thin bark, that is trees smaller than 16 cm dbh. A second wasp, Spathius agrili, doesn’t persist in northern regions – i.e., northern Minnesota!  Therefore they are seeking additional management tactics workable in natural mature ash stands across northern North America.

Peters, Rajtar, & Blanchette (2025) are testing whether useful levels of suppression can be provided by release of entomopathogenic fungi (EPF). They are building on earlier studies link? that determined that one entomopathogenic fungus isolate, Beauveria bassiana, showed promise in reducing EAB population growth. This fungus is available as a commercial product.

Peters, Rajtar, & Blanchette (2025) tested isolates from five genera = Beauveria, Purpureocillium, Metarhizium, Clonostachys, & Samsoniella. They found that Beauveria pseudobassiana, Beauveria bassiana GHA, Metarhizium sp. Meta, & Purpureocillium sp. consistently reduced EAB mean survival time (MST) and probability of survival over time.

They also tested use of an inoculation method different from that used in the earlier studies:  autodissemination devices (ADDs). The devices expose adult EAB to the fungal spores using baited Lindgren funnels with non-stick surfaces hung in the canopy of an infested ash tree. Infected beetles transmit the fungus to others during mating. In a Canadian study, this method of infecting EAB adults with an isolate of B. bassiana CFL-A caused rates of infection up to 40% and a significant decrease in EAB pop growth rate over a 3-year period.

Scientists have isolated and identified 1,126 fungal isolates associated with EAB larval galleries from trees growing across Minnesota. Eight percent of these isolates are entomopathogens. Some of these isolates have been proved to kill EAB eggs. In this study Peters, Rajtar, & Blanchette (2025) tested the fungi’s efficacy against adults and whether the ADDs system inoculates sufficient numbers of beetles.

They found that five of the fungal isolates significantly decreased the mean survival time of EAB adults post-exposure and the probability of EAB survival over time. These were Beauveria bassiana GHA, B. pseudobassiana EAB 16.8 & B. pseudobassiana EAB 53-5, Metarhizium sp. Meta, and Purpureocillium sp. EAB 59-16-2. They are most promising isolates for future study. Peters, Rajtar, & Blanchette (2025) discourage further consideration of isolates in the genera Samsoniella and Clonostachys because of disappointing reductions in adult beetle survival.

The probability of survival of beetles in the negative control groups differed significantly between the two experiments. The authors say they can only speculate about the reason because many external factors influence EAB adult lifespan. Among such factors are a) the longer storage period for insects used in the first experiment; b) differences in nutrient quality or microbiomes of ash leaves provided during incubation; or c) health of trees from which the beetles had been collected.

Peters, Rajtar, & Blanchette (2025) say the effectiveness of these fungi and this inoculation method must now be tested under field conditions. One concern: the fungal inoculum can degrade due to exposure to the environment while it is in the trap. A second concern is that the fungus might grow across the inoculum pouch unevenly, thereby undermining equal delivery of conidia to each beetle. They call for  research to optimize the type and/or formulation of inoculum used in ADDs to prolong fungal viability in the field.

an ash in a suburban setting; photo by F.T. Campbell

While I am cheered by the prospect of conservationists having new tools to counter the EAB infestation, I am concerned by how slowly work pursuing possible use of entomopathogens is proceeding. Peters, Rajtar, & Blanchette (2025) refer to several past studies that reached similar findings. Why has this knowledge not been applied in the field already? Is it due to the intrinsic difficulty of detecting and working with the fungi? Lack of funding for non-commercial approaches? Leaders’ narrow focus on strategies already in use? Some other complication?

SOURCES

Peters, C.J.; Rajtar, N.N.; Blanchette, R.A. Entomopathogenic Fungi from MN Are Virulent Against Emerald Ash Borer, Agrilus planipennis Fairmaire (Coleoptera: Buprestidae), Adults in a Laboratory Autodissemination Device Assay. Forests 2025, 16, 1742. https:// doi.org/10.3390/f16111742

Trotter, R.T., III; Warden, M.L.; Vazquez, R.J.; Ryan, J.K.; Pfister, S. 2024. ALB Hazard Management and Monitoring Version 1.0: an assessment and tracking tool for Asian longhorned beetle eradications in the United States. Gen. Tech. Rep. NRS-222. Madison, WI: U.S. Department of Agriculture, Forest Service, Northern Research Station. 24 p. https://doi.org/10.2737/NRS-GTR-222.

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

Or

https://fadingforests.org

Port of Houston expands – is the region ready for incoming wood-borers?

Barbours Cut terminal, Port of Houston; photo by Patrick Feller via Flickr

I blogged recently about ports in the southeast dredging channels and taking other actions to increase the number of container ships they can service. Michael Angell of the Journal of Commerce reports that the Port of Houston is competing vigorously for visits by ships from Asia. Of course, rising ship visits increase the opportunity for Asian pests such as Asian longhorned beetle or several ambrosia beetles in the Euwallaceae complex to invade the Gulf Coast region. (The emerald ash borer is already present in Texas.)

Considering the immediate future, the Port of Houston is expanding capacity of its two existing terminals. The Port is dredging portions of its channel down to a 46-foot depth. It expects to complete dredging of the main channel to 55 feet – the depth of Norfolk currently by ten years from now.

One of the existing terminals, the Bayport container terminal, handles most of the trans-Pacific Asia services coming into the Gulf Coast. Two new berths and storage yard space are being added, so that Bayport will be able to handle five super-post-Panamax ships simultaneously. The other terminal, Barbours Cut terminal, is also being renovated to improve handling of super-post-Panamax ships. At full buildout, the Bayport and Barbours Cut will have capacity for handling 8 million TEUs per year.

Houston has begun design work for a third container terminal. At full capacity – more than a decade from now — this third terminal will handle another 4 million TEUs. The project has been approved by the Federal Permitting Improvement Steering Council for streamlined environmental under the 2015 FAST Act. This program does not alter any applicable statutory or regulatory requirement, or guarantee approval. It seeks to expedite lawfully mandated environmental and other review through standardized interagency consultation and coordination practices. It also created a public online tool to track progress in completing these requirements. 

Through May 2026, Houston handled 502,387 TEUs in container imports from Asia, up 10% from the same period in 2025. The number of container ships over 10,000 TEUs in capacity calling at Houston doubled, from four in the second half of 2025 to nine during the first half of 2026.

Port of Long Beach

Houston ranked fifth in container volumes during this period. Los Angeles was the top port, with 1,974,330, followed closely by Long Beach with 1,851,159. Together, the two California ports received more than seven times as many containers as Houston. Houston is competing with other ports in the Southeast. Savannah still handles more than a third higher numbers than Houston, at 823,823 containers. The Port of Virginia (see earlier blog) received 285,046 containers, and Charleston only 256,499 Both were decreases from the 2025 number. All three southeastern ports saw small decreases from their 2025 numbers.

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

Or

https://fadingforests.org

Florida’s bromeliads: who will protect them?

Tillandsia utriculata; photo courtesy of Teresa Cooper

Florida has 16 native species of bromeliads. Up to 15 invertebrates depend on bromeliads, especially the water that collects at the bases of the leaves. See Frank (1983) and Frank and Fish (2008) for more information about bromeliads’ ecological importance.

Eight species had been listed as threatened or endangered in the state (visit here)  (Frank and Cave 2005) before arrival of the Mexican bromeliad weevil. The Florida Endangered Plant Advisory Council added two of its hosts – Tillandsia utriculata and Tillandsia fasciculata. The website now lists 12 species as threatened by the weevil.

Twelve species are believed to be vulnerable to the weevil. T. utriculata is at most immediate risk. Before arrival of the Mexican bromeliad weevil, some populations of this species were described by Teresa Cooper as “giant” and dense, containing very large mature plants and producing huge volumes of of seed.

One such population — in the Enchanted Forest Sanctuary in Brevard County – succumbed to the weevil quickly: 87% died in six months. At 27 months after arrival of the weevil, less than 3% of the original population remained. It is now rare to see large T. utriculata plants. Medium-sized plants put out inflorescences. The weevil persists at low levels, so is able to feed on and reproduce on medium and small plants. Several endemic wildlife species that inhabit the water pooled in T. utriculata’s base face global extinction (T.M. Cooper pers. comm. 2026).

damage to a bromeliad caused by the weevil; photo courtesy of Teresa Cooper

The first reason T. utriculata is at higher risk is that each plant can support many weevil larvae. Its leaves are softer (so more easily consumed) and its large stems contain lots of nutrients — for a bromeliad). More important is the plant’s very slow reproduction in contrast to the weevil’s rapid attack. T. utriculata plants do not produce seed until after they have grown for up to 20 years. Furthermore, an extremely small fraction of the ~ 10,000 seeds released by a mature plant make it to reproductive age. A plant that after 18 years begins growing an inflorescence — which then takes ~ a year to flower and produce seed – can be killed by a weevil within a few months. In addition, T. utriculata is monocarpic – it doesn’t produce offshoots. T.M. Cooper asserts that if the bromeliad weevil extirpates T. utriculata from Florida, it will not be because the weevil ate the very last plant, it will be because T. utriculata seed output has fallen below a sustainable level.

Teresa Cooper demonstrating the size of a T. utricuclata flower spike; photo courtesy of T. Cooper

A second host, Tillandsia fasciculata, had large and dense population in Loxahatchee National Wildlife Refuge (Palm Beach County) in 2002 to 2005. Almost every tree had multiple T. fasciculata plants; many were huge with multiple pups. When Teresa Cooper revisited about 10 years later, T. fasciculata plants were very sparse, large plants very rare; none was really large (T.M. Cooper pers. comm. 2026).

T. fasciculata is expected to decline more slowly than T. utriculata. First among several advantages is that its leaves are tougher and stems smaller, so they provide less nutrition to the weevil. Probably more important, T. fasciculata is polycarpic (it produces both seed and offshoots). A large T. fasciculata plant can sprout up to a dozen blooming rametes at a time. If the weevil kills one, other rametes will bloom and seed. At the same time, the plant might produce numerous up to 20 offshoots which come detatched from the “mother” plant and start a new individual. Still, as the weevil kills these rametes year after year the plant shrinks in size and eventually dies (T.M. Cooper pers. comm. 2026).

Teresa Cooper has documented damage on Tillandsia paucifolia, T. simulate, and T. variabilis (T.M. Cooper pers. comm. 2026). She has no funding to continue studying these species. I have been unable to find any more recent information about the status of the Tillandsia species. The Florida Department of Agriculture and Consumer Services does not have an active program addressing the weevil (Patrick Gordon, FDACS, pers. comm. June 2026).

There is considerably more information about another endangered bromeliad, Guzmania monostachia. This species has the broadest distribution of any species in the genus, stretching across northern South America, Central America, the Caribbean, as well as southern Florida. As the northernmost population, the Florida population might harbor unique genetics important for the species’ adaptation to climate change (Krupar et al. 2023).

Formerly more common, G. monostachia is now found in six fragmented and disjunct populations in five areas in Florida. The species has apparently been extirpated from four other sites by habitat loss, not depredations by the weevil. One set of habitats is in wetland sloughs on the peninsula’s west coast on the peninsula’s west coast. These forests comprise pond apple (Annona glabra) and Florida ash (Fraxinus caroliniana); Kupar et al. (2023) don’t mention whether this species is vulnerable to the emerald ash borer, which has not yet been detected in Florida. The largest of the bromeliad populations is in Fakahatchee Strand State Preserve, in Collier County in Southwest Florida. Fakahatchee Strand Preserve is Florida’ oldest and largest state park.  The bromeliad’s population was previously estimated to exceed two million individuals. Their number has been halved by weevil attack. Also, since M. callizona prefers larger plants, the proportion of reproductively mature individuals had been reduced from roughly 50% to only 10–20% by 2021. T.M. Cooper reports (pers. comm. 2026) fears this species might also be extirpated eventually.

Guzmania monostachia; photo courtesy of Teresa Cooper

The adjacent Big Cypress National Preserve is home to two populations of G. monostachia; they are separated by ~2,900 km2 so there is no interaction between them. The southern population consists of ~ 1000 individuals, the northern population comprises only~ 200 individuals.

Guzmania monostachia is in an even more precarious situation on the eastern side of the peninsula: tiny populations of two or three plants are found in three locations: Everglades National Park, Fuchs (formerly Sykes) Hammock Preserve, and Meissner Hammock Preserve.

The Mexican bromeliad weevil Metamasius callizona (Chevrolat) is native to southern Mexico, Guatamala, (Frank and Thomas 1994; Frank and Cave 2005) and Belize (Cooper, Cave and Frank 2023). Its presence in Florida was detected in a nursery in Ft. Lauderdale, Broward County, in 1989. Probably introduced on a shipment of ornamental bromeliads from Mexico. The nursery treated the infested plants, but the weevil had already established on native bromeliad species in the natural environment. By 1991, it was detected in four counties in southern Florida; by 1999, it was found in 12 more. It is now in most counties of peninsular Florida from Daytona (Volusa County) south to Miami-Dade (Patrick Gordon, Florida Department of Agriculture and Consumer Services, pers. comm. June 2026).

While chemical control is feasible in nurseries and display plantings, it can’t be used in natural areas, where the epiphytes are not accessible from the ground and stakeholders fear likely non-target effects.

Therefore, scientists initially focused on classical biocontrol. They made 16 expeditions to Mexico and Central and South America from 1992 to 2010 searching for natural enemies of the weevil. The chose parasitoid tachinid fly, Lixadmontia franki. A colony was established in the University of Florida’s quarantine laboratory and research on fly-weevil interactions was conducted for several years. Releases into the environment were begun under Federal and State permits in 2007. Although more than 3,100 flies were released, the species failed to establish. So this approach is no longer being pursued (Cooper, Cave, and Frank (2024); T.M. Cooper pers. comm. June 2026).

Meanwhile, in 2010, Dr. Frank and D. Giardina of the Florida Fish and Wildlife Conservation Commission observed that in one location in Belize the weevils had no apparent detrimental effect on bromeliads identified as Tillandsia utriculata (although there is some disagreement on this classification). Cooper, Cave, and Frank (2024) undertook a study to determine how the bromeliad could coexist with Metamasius callizona in Belize while being so vulnerable in Florida. They compared life cycle parameters (oviposition and pupation rates, egg hatch rate, adult emergence and size, and developmental time) of weevils from Florida and Belize. They also observed how populations of the weevil from Florida fared on three hosts: pineapple (Ananas comosus), T. utriculata collected from Florida, and T. utriculata collected from Belize. Finally, they quantified the hosts’ nutritional content using two measures (percent soluble solids and leaf toughness). 

Their most important finding is that weevil larva from Florida could not develop past the third instar on leaves of the T. utriculata from Belize. The weevil larvae starved because the Belize bromeliad’s leaves had fewer nutrients and were tougher than the leaves from T. utriculata collected in Florida.

bromeliads seen in Big Cypress National Preserve in the early 1980s; photo by F.T. Campbell

Based on this finding, Cooper, Cave, and Frank (2024) recommend that authorities introduce T. utriculata plants from Belize into Florida’s forests and allow them to colonize and/or hybridize naturally with Florida’s remaining plants. An alternative would be to breed hybrid Tillandsia in the lab and release them into Florida’s forests. I note that the American Chestnut Foundation pursued a similar strategy for decades to develop American chestnuts (Castanea dentata) able to tolerate the chestnut blight fungus (Cryphonectria parasitica).

Sources

Cave, R.D. 1997. Admontia sp., a potential biological control agent of Metamasius callizona. J Brom Soc. 47:244-249.

Cave, R.D. 2008. Biological control of the MEXICO bromeliad weevil. Biocontrol News and Information 29(1):1N-2N.

Cooper T.M. 2006. Ecological and demographic trends and patterns of Metamasius callizona (Chevrolat), an invasive bromeliad-eating weevil, and FLORIDA’s native bromeliads [Master’s thesis]. [Gainesville (FLORIDA)]: University of FLORIDA. 69 p.

Cooper, T.M., R.D. Cave, and J.H. Frank. 2023. Potential bottom-up control of Metamasius callizona in  Florida, USA. Entomologia Experimentales et Applicata. 2024. 172;4090421

Frank, JH. 1983. Bromeliad phytotelmata and their biota, especially mosquitoes. In: Frank J.H., Lounibos, L.P., editors. Phytotelmata: terrestrial plants as hosts for aquatic insect communities. Medford (NJ): Plexus. p. 101-128.

Frank, J.H. 1996. Bromeliad biota: history of Metamasius callizona FLORIDA [online]. Gainesville (FLORIDA): University of FLORIDA [cited 2010 Feb 1]. Available from https://www.entnemdept.uFlorida.edu/frank/bromeliadbiota/wvbrom6.htm

Frank, J.H. and R.D. Cave. 2005. Metamasius callizona is destroying FLORIDA’s native bromeliads. In: Hoddle MS, editor. USDA Forest Service Publication FHTET-2005-08. Vol 1. Second International Symposium on Biological

Control of Arthropods; 2005 Sep 12-16; Davos, Switzerland. Washington D.C.: USDA Forest Service. p. 91-101.

Frank, J.H. and D. Fish. 2008. Potential biodiversity loss in Florida bromeliad phytotelmata due to Metamasius callizona (Coleoptera: Dryophthoridae), an invasive species. Florida Entomol. 91(1):1-8.

Frank, J.H. and M.C. Thomas. 1994. Metamasius callizona (Chevrolat) (Coleoptera:Curculionidae), an immigrant pest, destroys bromeliads in FLORIDA. Can Entomol.126(1):673-682.

Krupar, S., A.A. Naranjo, G. Godden, N. Cellinese. The Fate of Guzmania monostachia in Florida Rests with Humans. Diversity 2023, 15, 525. https://doi.org/10.3390/d15040525

Potter, K.M., Escanferla, M.E., Jetton, R.M., Man, G., Crane, B.S., Prioritizing the conservation needs of US tree species: Evaluating vulnerability to forest insect and disease threats, Global Ecology and Conservation (2019), doi: https://doi.org/10.1016/

Salas, J. and J.H. Frank. 2001. Development of Metamasius callizona (Coleoptera:Curculionidae) on pineapple stems. Florida Entomol. 84(1):123-126.

Wood, D.M. and R.D. Cave. 2006. Description of a new genus and species of weevil parasitoid from Honduras (Diptera: Tachinidae). Florida Entomol. 89(2):239-24.

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

Or

https://fadingforests.org

Update on invasive plants in Hawai`i

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 Invasions 15, 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

Or

https://fadingforests.org

Invasive species on the African Continent

We are beginning to get more information about invasive species on the African continent.

a flyer naming principal invasive ornamental plants in Kruger National Park

In several countries, the focus has been on threats to agriculture. Previous blog about horizon-scanning in Ghana. In Zimbabwe, N. Mudada and colleagues (2026; full citation and the end of this blog) found alarming, if not surprising, levels of risk to food production from introduced invasive plants. They investigated 1,668 human-aided transboundary plant introductions at 14 ports of entry and non-official crossing points over the course of four years.

They estimate that the 20,000 trucks that transported maize into the country over the four years carried over 20,700 metric tons of weed seeds and rubbish! They recorded detections of 11 species in eight orders. The pathways are familiar. As noted, several weeds were contaminants of grain shipments; Convolvulus arvensis in wheat for human consumption, Helianthus annus and Datura stramonium in maize for animal feed. Adenium obesum and Vitex agnus-castus were being smuggled for planting as flowers and ornamentals. (Vitex agnus-castus was also smuggled in passenger baggage for its medicinal properties). Several Lemna species (an aquatic plant) were also smuggled for planting as animal feeds.

In some cases, the focus is the threat to native ecosystems. I posted a blog the about threat of an introduced pathogen to trees in the remnant rain forests of Madagascar.

South Africa still has the lead in addressing invasive species. Regarding invasive plants specifically, the country has the benefit of more than 150 years of botanizing. The richness of the region’s flora is globally recognized. South Africa also has a long history of studying and managing invasive species, especially plants.

South African scientists and colleagues in Botswana, Eswatini, Lesotho, and Namibia have published four editions of the Flora of the Southern Africa region since 1984. In 2006, the PRECIS database of the South African National Biological Diversity Institute (SANBI) was combined with the Tropical African Plant Checklist published by the Conservatory and Botanical Garden of Geneva to create the African Plant Checklist and Database Project. It is continually updated. This is the first continental flora checklist for Africa; it fulfils countries’ obligations under the Convention on Biological Diversity’s Global Strategy for Plant Conservation.

For South Africa specifically, scientists have produced a national plant checklist that is updated annually.

The 2025 Checklist reports that 21,539 plant species are extant outside cultivation in the country; these comprise 20,204 indigenous species and 1,329 naturalized species. Thus, 6% of the total flora is non-indigenous. Of these, 649 (48.8% of the non-indigenous species, 3% of all plants) of them are invasive.

[Naturalized species are defined as species whose documented natural range does not include South Africa, but have overcome a biogeographic barrier and now sustain self-replacing populations for two or more life cycles or over a given period of time in the country. These populations are maintained without direct intervention by people, or despite human intervention. Invasive species meet the above definition plus produce reproductive offspring, often in large #s at considerable distances from the parent and/or site of introduction, and have the potential to spread over long distances.]

Since the previous checklist was published in 2006, botanists have identified 1,048 additional species – a 4.9% increase. Eighty-two percent of the newly identified species (865 species) are “naturalized”. Specifically, 414 new species are categorized as naturalized (a 31.1% increase), and 451 new species are classified as invasive (a whopping 69.5% increase). Le Roux and Klopper attribute these steep increases to active botanizing by SANBI’s Invasive Species Programme (begun in 2008), and the Southern African Plant Invaders Atlas Project (begun in 2010).

Of the 384 plant families present in South Africa, 350 contain at least some indigenous species. Thirty-four families contain only naturalized species. Among the 2,189 plant genera present, 459 (21%) contain only species that are non-indigenous.

Three families stand out because of the particularly high numbers of naturalized species: Fabaceae (143 species; 11% of all naturalized species), Asteraceae (140 species; 10%), Poaceae (123 species; 9%). Two of these families — Asteraceae and Fabaceae — are also the largest families among native South African plants. The third, grasses (Poaceae), ranks seventh in the list of most specious families indigenous to South Africa. The next group of families with high numbers of naturalized species has less than half as many invasive species: Myrtacae (55), Amaranthacea (52), Solanaceae (48). None of these families ranked within the top 20 families of indigenous plant species.

The genera with the most naturalized species were Solanum, Euphorbia and Acacia (all 24 or 23 species).

Acacia cyclops; photo by David M. Richardson

South African scientists are also exploring how to balance conflicting goals and perspectives when an invasive plant species has economic or social value. The example chosen by Mbobo et al. (2025) is guava (Psidium guajava) – a nutritious and popular tropical fruit grown commercially in South Africa, but also invasive along roadsides, watercourses and forest margins. Invasions are especially common in eastern parts of country; large monocultures are found in KwaZulu-Natal. Outbreaks have also been detected at five sites in Western Cape in riparian zones and at a hot spring. Mbobo et al. (2025) note that the microclimatic conditions at this last location differ from the broader conditions in the region – which are what most models would measure.

The scientists used models to predict where guava might invade – especially in large monocultures – and compared those areas to where the tree can be grown in cultivation with human inputs, e.g., irrigation. They then assessed whether six regulatory approaches would avoid restricting guava farming in areas at minimal or low risk while still protecting vulnerable locations. They also considered the amount of information required to implement the approach and costs of acquiring the information; and level of likely public acceptance. Mbobo et al. (2025) laid out the trade-offs between continuing to regulate planting of the species at the provincial level vs. at the municipal level. Prohibiting planting of guava in provinces where it is recorded as invasive allows some plantings near natural forests and riparian areas that are highly susceptible to guava invasions. On the other hand, nearly half of the prohibited area is outside the known or likely at-risk area. The provinces do allow exceptions through a permit process. Adopting more geographically limited rules by regulating at the municipal level would enable a tighter link to geographic areas at highest risk. However, this approach does not address long-distance seed dispersal by animals. Furthermore, the very detailed regulations might confuse stakeholders and complicate enforcement. Also, the models lack sufficiently fine spatial resolutions to predict invasible areas so accurately. Finally, the reduction in regulated area is minimal (~ 14%), so the economic benefits are unlikely to outweigh the significantly higher administrative costs and risk of allowing guava invasions in new sites.

Guava fruit on tree; Roenashy via Wikimedia

Gildenhuys et al. (2026) analyzed the factors that drive which non-native plants establish where. They assessed the roles of temperature, precipitation, urbanization intensity, urban area, travel time, year of city’s establishment, and human population density in determining which plant species are present in 54 urban centers in Western Cape Province. The cities have significant differences in climate: Mediterranean in the far southwest, warm temperate in the southeast, and semi-arid towards the interior. The expectation was that these drivers and assembly processes are influential at more advanced invasion stages when the species have already overcome some barriers to dispersal, so are now found in reasonably suitable habitats.

Gildenhuys et al. (2026) found temperature and precipitation were most important in determining plant species’ presence. This was especially true at the boundary between arid and mesic climates. These strong environmental gradients are the same ones which have driven high differences in native species presence across the province. [See pamphlet describing invasive plants in Cape Town.] This finding supports the “Goldilocks hypothesis”: that non-native plant species assemblages are driven by the same abiotic variables as native species assemblages. While did not directly study the “Biome decides hypothesis” (the composition of non-native flora is mediated by the biotic effects of native flora and fauna), Gildenhuys et al. (2026) doubt its applicability here because native species’ presence has probably been greatly reduced by the effects of urbanization.

Urbanisation intensity itself ranked third as a factor. Its effect was strongest at low to medium urbanization intensities. Because urbanization creates novel habitats, such as, “hardscapes” of paved surfaces that resemble deserts, their non-native plant assemblages are dominated by similar, urban specialist, species. At lower urbanization intensities a greater variety of habitats is available. Gildenhuys et al. (2026) conclude that urbanization acts primarily as a driver of opportunistic habitats for species at later invasion stages rather than as a filter of species introduction.

An earlier study found a similar effect from road density (often associated w/ urbanization) as an explanation for where specific woody non-native species establish. They do concede that larger urban areas might experience greater propagule pressure.

Gildenhuys et al. (2026) note that recent globalization of the plant trade has probably changed the specis planted in urban centers. For example, cities in the Western Cape are increasingly replacing English oak (Quercus robur) with more disease-resistant oaks. The change might reflect greater environmental awareness and regulations issued under the National Environmental Management: Biodiversity Act 10 of 2004. In newly established urban centers, fewer invasive species are being planted — at least among trees.

SOURCES

Gildenhuys, C.P., L.J. Potgieter, C. Hui, D.M. Richardson. 2026. Drivers of compositional turnover of the NIS urban flora in the W Cape, South Africa. Urban Ecosystems (2026) 29:51 https://doi.org/10.1007/s11252-026-01919-3

Le Roux, M.M., R.R. Klopper. 2025. Taking stock of South Africa’s flora. South African Journal of Botany 184 (2025) 571-579

Mbobo, T., D.M. Richardson, A. Datta, K.T. Faulkner, J.R.U. 2025. Wilson. Spatially-Differentiated Reg of NIS Can Be Improved Using Spp Distribution Models: Psidium guajava in South Africa as a Case Study. Diversity and Distributions. 2025 31:e70102 https://doi.org/10.1111/ddi.70102

Mudada, N., J. Chitamba, E. Nyangani, C. Chapano, N. Mapope,and W. Ngezimana. 2026. Weeds associated with cross border traffic, their approach and infestation rates in Zimbabwe.  ISABB Journal of Food and Agricultural Sciences. Vol. 12(1) January-June 2026. DOI: 10.5897/ISABB-JFAS2025.0192

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

Or

https://fadingforests.org

Floral richness of South Africa

Cape Coast lily (Crinum macowanii); photo by Jaqui Geux

South Africa is rich in plant diversity. That diversity is strongly shaped by the country’s varied topography, soils, & climate. Nine biomes & 465 vegetation types are recognized in a country of just 1.2 million km2 (471,445 square miles).

There is a long history of botanizing in southern Africa – with impressive results that include a Flora of the Southern Africa region (it covers Botswana, Eswatini, Lesotho, Namibia, as well as South Africa) and the African Plant Checklist & Database Project, which is the first continental checklist for Africa. This publication – which is being maintained continuously – contributes to goals of the Convention on Biological Diversity’s Global Strategy for Plant Conservation.

South Africans have also produced a national plant checklist that is updated every year. The most recent Checklist dates from 2025. The checklist includes bryophytes (mosses, hornworts, & liverworts), lycophytes & pteridophytes, gymnosperms, & angiosperms. Compilers plan to include marine macro-algae in future.

The 2025 checklist names 21,539 species in South Africa – 20,204 indigenous species, 1,329 naturalised species. Based on the, South Africa’s national flora comprises 5.3 % of the estimated ~ 380,000 plant species on Earth.

Geissoloma marginatum; photo by Tony Rebelo via Wikimedia

South Africa is home to 384 plant families, of which all but 34 contain only indigenous species. All the species in seven families are endemic to South Africa. These families are Bruniaceae, Geissolomataceae, Grubbiaceae, Lanariaceae, Penaeaceae (excl. Oliniaceae), Rhynchocalycaceae, & Roridulaceae. Of the 2,654 genera, 2,189 are indigenous. Three hundred (12%) of these genera are endemic. Sixty percent of the 20,204 indigenous species are endemic. All the endemic families other than Bruniaceae are composed of a single genus – demonstrating the phylogenetic uniqueness of this flora.

The 2025 Checklist contains 1,048 species that did not appear in the previous checklist (published in 2006). This is an increase of 4.9%. Numbers of indigenous taxa increased by less than 5%: additional 623 indigenous species (3.1%); 520 endemic species (4.3%). Numbers of non-indigenous species increased by considerably higher proportions: naturalized species increased by 414 species (31.1%), & invasive species by 451 (69.5%). These 865 species constitute 82.5% of all 1,048 newly recorded species. (I report the findings on non-indigenous species in a separate blog.)

The detection of previously unidentified species and infraspecific taxa – both native and non-indigenous – is the result of systemic botanizing campaigns focused on particular families. Thus, a study of the family Iridaceae in southern Africa led to description of 169 species in 20 genera. More than two decades of work on the Pelargonium (Geraniaceae) resulted in recognition of 34 new taxa. Thirty-three new species have been described in the genus Indigofera (Fabaceae).

The most specious plant family indigenous to South Africa is Asteraceae, with 2,124 species. Aizoaceae & Fabaceae follow with 1,603 & 1,566 species, respectively. The newly expanded Iridaceae is fourth; it encompasses 1,189 species. The world-famous Proteaceae rank 13th, with only 355 species. Considering genera, the largest by far is Erica (heath) at 734 species; no other genus houses more than 300 species. [The famous orchid genus Disa ranks 18th. There are ~182 species on the continent, primarily in East & Southern Africa.]

Disa uniflora on Table Mountain; F.T. Campbell

When considering phylogenies with high levels of endemism, the leading families are – again — Aizoaceae (1,426) and Asteraceae (1,352). Ranked third is Iridaceae (983). Fabaceae is ranked 4th (961). Nearly half (49%) of species in 1,071 genera are endemic. Again Erica ranks highest: 581 species – 78% of the species in the genus – are endemic. This is double the number in the genus ranked second — Aspalanthus (261species; all but 34 species are endemic). Pelargonium is 3rd : 204 of 261 species are endemic.

Flora of the provinces

 Of South Africa’s nine provinces, KwaZulu-Natal (94,361 km2) is home to the greatest diversity in terms of plant families (331 families) and genera (1,718 genera). Neighboring Eastern Cape Province (168,966 km2) ranks second with 305 families and 1,576 genera. Western Cape has somewhat fewer plant families (294) and genera (1,475), but by far the most species (11,379). Western Cape also leads in the proportion of its plant taxa that are endemic to the province: 58%. (This reflects the presence of a globally recognized distinct flora, the Cape Floral Kingdom.) KwaZulu-Natal and Eastern Cape (16% endemic) share most of their plant taxa with neighboring provinces or countries – Mozambique. Therefore, their levels of endemism are only 8% and 16%, respectively.

The clustering plant diversity and endemism in the south & east reflects the topographic variation provided by the Escarpment of South Africa.

The largest province is Northern Cape Province at 372,889 km2. It is home to 5,058 species but its fairly uniform terrain means there is not very high species diversity (25 % – half the proportion in Western Cape). However, the flora is unique because adapted to the harsh environment (I concede these statements seem contradictory).

Pachypodium namaquanam; F.T. Campbell

SOURCE

Le Roux, M.M., R.R. Klopper. 2025. Taking stock of South Africa’s flora. South African Journal of Botany 184 (2025) 571-579

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

Or

https://fadingforests.org

EAB infestation at 20 years: focus on green & white ash & white fringetree

dying ash in Shenandoah National Park; photo by F.T. Campbell

The emerald ash borer (Agrilus plannipennis; EAB) was detected in North America in 2002. So both U.S. and Canadians have been motivated to evaluate the probable trajectory of the primary hosts – the ash genus Fraxinus – in the face of the ongoing invasion. See Deschênes et al. 2026 and Wilson et al. 2025 – full citations at the end of this blog. Both studies focused on white (Fraxinus americana) and green ash (F. pennsylvanica); they say next to nothing about black ash (F. nigra). I regret this silence because of the unique ecology of black ash swamps. Neither addresses the threat to Oregon ash (F. latifolia) in the West.

The two assessments have similar findings: high mortality of larger trees (canopy and “recruit” size trees); abundant regeneration (seedling and saplings sizes) after an initial period; and uncertainty as to whether persisting EAB populations will kill the saplings before sufficient numbers grow into reproductive size.

There are two conspicuous differences. First, the American study does not consider the possible impact of biological control – although USDA APHIS has placed all its effort on this approach since January 2021. The Canadians report that self-sustaining populations of the wasps are now found across the northern U.S. and eastern Canada. In their study, conducted in Ontario, they detected only Tetrasticus planipennisi; it was parasitizing 16% of the EAB larvae in dissected stems. This wasp’s affinity for colder climates and short ovipositor – which limits it to parasitizing larvae inhabiting small stems – are portrayed as positive traits under these circumstances.

Second, the Canadians did not find “lingering” adult ash trees as have the Americans. These trees indicate the probability of finding workable levels of genetic resistance to the EAB. USDA Forest Service scientists are pursuing a breeding program.  While in south-central Michigan, where overstory ash mortality typically exceeded 80%, 46% of overstory ash and 82% of ash recruits were relatively healthy (Wilson et al. 2025), in Ontario none of 1,129 overstory ash trees survived beyond seven years after EAB detected. No trees exceeded 15 cm dbh (Deschênes et al. 2026).

ash resistance breeding trial at Holden Arboretum; photo courtesy of Jennifer Koch, USFS

Regeneration

As Deschênes et al. (2026) state, the future of ash stands depends on the complex of interactions among environmental conditions, management interventions, efficacy of natural enemies (natural or introduced), and life-history traits of the insect and its host. Coexistence might be possible if EAB larval densities remain sufficiently low to support survival of residual trees and successful seedling recruitment.

Larval densities in Ontario were said to be generally low, suggesting reduced carrying capacity in post-invasion forests, lower EAB fecundity, and higher EAB mortality in regenerating stems. Deschênes et al. (2026) note that in more northern areas colder temperatures are thought to slow larvae development. Perhaps these larvae might also be less vigorous, so they night disperse only over short distances. Still, there were sufficient EAB present after all the overstory trees had died to create 97% of the 298 galleries in regenerating ash stems (Deschênes et al. 2026). Furthermore, Wilson et al. (2025) say that EAB densities in infested trees in Michigan were similar to densities recorded during the initial invasion. This seems ominous to me – a solid foundation from which beetle populations could build up again as regenerating ash grow and provide more phloem for the insect to exploit.

Ash reproduce by both flowering/seeding and sprouting from the base. EAB predation is not the only complication. First, ash are dioecious so mature trees of both sexes must grow within a few hundred meters. Second, predation by the ash seed weevil (Lignyoodes helvolus) reduces seed supplies. Dense sedge mats can prevent germination (Wilson et al. 2025). Scientists generally believe that the soil seed bank is quickly exhausted, although Wilson et al. (2025) cite others’ conflicting findings. Neither article discusses predation by mammals, e.g., deer or rabbits. Wilson et al. (2025) mention attacks by beavers.

ash saplings felled by beavers; photo by F.T. Campbell

Wilson et al. (2025) did not study whether stump-sprouted ash were able to successfully recruit into the overstory. They do report that in one study in southeastern Michigan stump sprouts were the dominant form of green ash regeneration and about a quarter of these sprouts produced seeds at least once. Deschênes et al. (2026) found that on average 47% of regenerating stems at their Ontario research sites originated from stump sprouts.

EAB has been documented to attack and kill trees when the main stem is as small as 2.5 cm. While EAB probably prefer larger stems, Deschênes et al. (2026) suggest that stems become acceptable at the lower range of size required for seed production – 8–10 cm dbh. Reliable and abundant seed production doesn’t occur until white or green ash achieve > 20 cm dbh. At their Ontario sites, Deschênes et al. (2026) found that 42% of regenerating stems has been infested by EAB at least once; 14% had been attacked five or more times. They removed 74 EAB larvae from 28 stems; 49 (66%) were alive. Fifteen EAB (16% of current year galleries) had been parasitized — all by Tetrastichus planipennisi. They also observed numerous signs of defensive responses.

In Michigan, no ash recruits — living or dead – were found in plots in 28% of the cells. In the remaining 128 cells, an average of ~33% of ash recruits were infested by EAB, and ~21.4% of ash recruits dead. As is typical, white ash fared better than green ash. Recruit sized ash trees were twice as likely to die than to survive and mature into overstory size (Wilson et al. (2025).

In Ontario, as noted, all canopy ash had died. There were 119 live trees 5 – 10 cm dbh – a tenth as many “mature” ash as when EAB arrived, and all were smaller. There was abundant regeneration in most sites initially, but at the longest-infested sites in Essex County, regenerating ash stems were half as numerous as early after the transition (Deschênes et al. 2026).

The Canadians found it encouraging that some of the regenerating stems were vigorous despite containing EAB gallery densities greater than 20 larvae·m?2 of phloem. They did not know the mechanisms underlying survival of these stems. Possible explanations ranged from the low EAB carrying capacity of smaller trees to stronger host defenses in regenerating stems to EAB mortality due to parasitism.

Wilson et al. (2025) note that despite more than 20 years of EAB presence, densities of ash recruits, saplings, and seedlings were high relative to other species. However, they remind us, ~ one-third of the live ash recruits were infested so their survival into reproductive size was uncertain. The high mortality of overstory ash results in loss of seed resources, greater sun exposure, and cascading consequences for forest composition and function. In upland sites, cells with low ash basal area favored Quercus rubra and Tilia americana. They conclude that changes to forest composition is probably site specific — largely depend on what tree species are already present.

Despite the challenges described above, the Canadian scientists also believe that these findings demonstrate that ash has a capacity for long-term regeneration (Deschênes et al. 2026).

Changing Species Composition in the U.S.  (Wilson et al. 2025)

Canopy gaps caused by ash mortality have largely been filled by lateral ingrowth of species already there — American elm (Ulmus americana), black cherry (Prunus serotina), and northern red oak (Quercus rubra). The regeneration strata (saplings and seedlings) is dominated by Fraxinus (white outnumbering green when differentiated), maples (Acer rubrum, A. saccharum), black cherry, Crataegus species and Carya ovata. Elms are consistently among most common non-ash taxa among overstory, recruit, sapling and seedling strata. At some Ohio sites there was also increased abundance of non-native tree and shrub seedlings. This is not surprising since invasive plants are widespread in the forests of Ohio and other eastern states. A decade ago 93% of Forest Inventory and Analysis (FIA) plots in Ohio had at least one of 50 invasive plant species.

In another paragraph they mention Tilia americana as one of the important species in these forests.

Situation in Canada (Deschênes et al. 2026)

Deschênes et al. (2026) express concern that the death of nearly all canopy-level trees will substantially reduce ash’ ability to fulfill its ecological roles in these ecosystems. Still, ash regeneration is persisting for decades following overstory mortality. The taxon’s continued presence is driven largely by strong sprouting, which has been observed in several locations in Ontario. In some areas, low EAB infestation rates and evidence that regenerating stems can withstand multiple infestations raises hope that some might reach maturity and produce seeds. This scenario would be similar to that of elms, in which surviving trees contribute to ongoing regeneration and might eventually facilitate development of some level of resistance to the invasive fungus. A second possibility is that ash’ high sprouting capacity might point to a scenario similar to that of American chestnut. This species has persisted for a century primarily as sprouting shrubs — although they rarely reach reproductive maturity.

white fringetree; photo by Ryan Somma via Wikimedia

White Fringetree

Scientists also reviewed the status of a secondary host of EAB in North America, white fringetree (Chionanthus virginicus). Earlier studies of this host-pest relationship had been conducted on ornamental plantings where the trees tend to be scattered across open lawns and actively managed – including protection from pests. The Cipollinis (see full citation at the end of this blog) believe they might be better able to ward of EAB attack than are wild, unmanaged trees in forests that must compete for resources. They wanted to assess the current status and likely trajectory of the tree species in the wild.

To do so they revisited a wild population of the tree in southern Ohio previously assessed 10 years earlier. White fringetree is a small multi-stemmed tree native to the southeastern U.S. It is widely planted as an ornamental in across the east. In Ohio, white fringetree grows wild in only a few southern counties, in small populations or as widely scattered individuals. The species is classified as “Potentially Threatened” at the state level.

In 2015, 30% of the white fringetrees at the site were infested. These trees had signs of stress but none had died. EAB larvae grow more slowly on fringetree than on North American ash species. Meanwhile, all mature white ash trees at the site had been killed by EAB. Smaller white ash trees more comparable in size to the white fringetrees had attack rates and impacts comparable to those on the fringetrees.

In their new study, the Cipollinis found that nine of 31 trees tagged in 2015 (29%) had died; 22 (71%) were alive. Of those 22 living trees, 12 (55%) stayed the same or improved slightly over the five-year period; 10 (45%) declined. Five of these 22 living trees (23%) had evidence of current infestation. Trees that had died had a higher incidence of old EAB galleries, adult exit holes, and woodpecker activity. This is interpreted as demonstrating that EAB must cause extensive damage to kill fringetrees.

In summary, fringetrees in a wild unmanaged population continued to be attacked by EAB over 10 years and suffered higher attack rates and more significant impacts than those previously observed in managed pops. The Cipollinis conclude that trees large enough to attract EAB oviposition will continue to decline in health and be killed as long as beetles are present. They expect that wild white fringetrees might meet the same fate as ash trees, but over longer time scales. 

At the same time, this delay in complete mortality might create a refugium for remnant populations of EAB after most ash have been killed. This status would be exacerbated if it turns out that the biocontrol agents cannot find their target — EAB — in the alternative host. The Cipollinis found lower parasitism rates by Tetrastichus planipennisi in fringetrees, although this was not true for the egg parasite Oobius agrili and two Spathius larval parasites.

Whitebark fringetree populations can produce few adult EAB because the trees are small and contain low amounts of phloem. Still, as young trees grow into vulnerable sizes they might help sustain the EAB population – as young ash trees in the area appear to do.

While caution is appropriate in interpreting findings from a study of a single population, the Cipollinis argue that this population has been studied intensively: assessed six times over 10 years, beginning at the start of the EAB infestation. Therefore they think their analysis provides useful informative regarding the long-term impacts of EAB on fringetree.

They concede that larger populations in areas deep within the tree’s native range might experience different dynamics and impacts. So far, however, observations in Chattahoochee National Forest in Georgia and at Great Falls Park on the Maryland-Virginia border generally support their finding that wild fringetrees in natural landscapes will suffer higher attack rates and be more severely impacted by EAB than trees in managed landscapes.

Finally, the Cipollinis fear that a close relative, pygmy fringetree, Chionanthus pygmaeus, is at particularly high risk because it is endemic to only a few counties in the sandhills of central Florida. The species is already classified as endangered by both the state and the federal governments. The pygmy fringetree is smaller than white fringetree, so its size might help it escape attack. However, adults achieve sizes comparable to that of fringetree in some cases. So when EAB reaches Florida, the specie appears to be highly vulnerable.

SOURCES

Cipollini, D. and K. Cipollini. 2026  The Fate of a Wild White Fringetree (Chionanthus virginicus) Population in Ohio 10 Years After Invasion by Emerald Ash Borer (Agrilus planipennis) Forests 2026, 17, 712

Deschênes, É., C.J.K. MacQuarrie, L. Scott, C. Zimmerman, and I. Aubin. 2026. Ash population dynamics after two decades of emerald ash borer infestations in Canada. Canadian Journal of Forest Research. Can. J. For.Res. 56: 1–13 (2026) | dx.doi.org/10.1139/cjfr-2026-0075

Wilson, C.J, L. Labbate, T.R. Petrice, T.M. Poland, D.G. McCullough. 2025. Ongoing regeneration of ash and co-occurring species 20 years following invasion by emerald ash borer. Forest Ecology and Management 580 (2025) 122546

 

Posted by Faith Campbell

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  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

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https://fadingforests.org