Category: invasive species

Global Overview of Bioinvasion in Forests

black locust – one of the most widespread invasive tree species on Earth; photo via Flickr

In recent years there has been an encouraging effort to examine bioinvasions writ large see earlier blogs re: costs of invasive species – here and here. One of these products is the Routledge Handbook of Biosecurity and Invasive Species (full citation at end of this blog). I have seen only the chapter on bioinvasion in forest ecosystems written by Sitzia et al. While they describe this situation around the globe, their examples are mostly from Europe.

Similar to other overviews, this article re-states the widely-accepted attribution of rising numbers of species introductions to globalization, especially trade. In so doing, Sitzia et al. assert that the solution is not to curtail trade and movement of people, but to improve scientific knowledge with the goal of strengthening biosecurity and control programs. As readers of this blog know, I have long advocated more aggressive application of stronger restrictions on the most high-risk pathways. Still, I applaud efforts to apply science to risk assessment.

Sitzia et al. attempt to provide a global perspective. They remind readers that all major forest ecosystems of Earth are undergoing significant change as a result of conversion to different land-uses; invasion by a wide range of non-native introduced species—including plants, insects, and mammals; and climate change. These change agents act individually and synergistically. Sitzia et al. give greater emphasis than other writers to managing the tree component of forests. They explain this focus by asserting that forest management could be either the major disturbance favoring spread of non-native species or, conversely, the only way to prevent further invasions. They explore these relationships with the goal of improving conservation of forest habitats.

Japanese stiltgrass invasion; photo by mightyjoepye via Flickr

Sitzia et al. focus first on plant invasions. They contend that – contrary to some expectations – plants can invade even dense forests despite competition for resources. They cite a recent assessment by Rejmánek & Richardson that identified 434 tree species that are invasive around Earth. Many of these species are from Asia, South America, Europe, and Australia. These non-native trees can drive not only changes in composition but also in conservation trajectories in natural forests. However, the example they cite, Japanese stilt grass (Microstegium vimineum) in the United States, is not a tree! Sitzia et al. note that in other cases it is difficult to separate the impacts of management decisions, native competitive species, and non-native species.

Sitzia et al. note that plant invasions might have a wide array of ecological impacts on forests. They attempt to distinguish between

  • “drivers” of environmental change – including those with such powerful effects that they call them “transformers”;  
  • “passengers” whose invasions are facilitated by other changes in ecosystem properties; and
  • “backseat drivers” that benefit from changes to ecosystem processes or properties and cause additional changes to native plant communities.

An example of the last is black locust (Robinia pseudoacacia). This North American tree has naturalized on all continents. It is a good example of the management complexities raised by conflicting views of an invasive species’ value, since it is used for timber, firewood, and honey production.

Sitzia et al. then consider invasions by plant pathogens. They say that these invasions are one of the main causes of decline or extirpations in tree populations. I applaud their explicit recognition that even when a host is not driven to extinction, the strong and sudden reduction in tree numbers produces significant changes in the impacted ecosystems.

American chestnut – not extinct but ecological role gone; photo by F.T. Campbell

Sitzia et al. contend that social and economic factors determine the likelihood of a species’ transportation and introduction. Specifically, global trade in plants for planting is widely recognized as being responsible for the majority of introductions. Introductions via this pathway are difficult to regulate because of the economic importance (and political clout) of the ornamental plants industry, large volumes of plants traded, rapid changes in varieties available, and multiple origins of trade. As noted above, the authors seek to resolve these challenges by improving the scientific knowledge guiding biosecurity and control programs. In the case of plant pathogens, they suggest adopting innovative molecular techniques to improve interception efficiency, esp. in the case of latent fungi in asymptomatic plants.

The likelihood that a pathogen transported to a new region will establish is determined by biogeographic and ecological factors. Like other recent studies, Sitzia et al. attempt to identify important factors. They name a large and confusing combination of pathogen- and host-specific traits and ecosystem conditions. These include the fungus’ virulence, host specificity, and modes of action, reproduction, and dispersal, as well as the host’s abundance, demography, and phytosociology. A key attribute is the non-native fungus’ ability to exploit micro-organism-insect interactions in the introduced range. (A separate study by Raffa et al. listed Dutch elm disease as an example of this phenomenon.)  I find it interesting that they also say that pathogens that attack both ornamental and forest trees spread faster. They do not discuss why this might be so. I suggest a possible explanation: the ornamental hosts are probably shipped over wide areas by the plant trade.

surviving elms in an urban environment; photo by F.T. Campbell

Sitzia et al. devote considerable attention to bioinvasions that involve symbiotic relationships between bark and ambrosia beetles and their associated fungi. These beetles are highly invasive and present high ecological risk in forest ecosystems. Since ambrosia beetle larvae feed on symbiotic fungi carried on and farmed by the adults inside the host trees, they are often polyphagous. Bark beetles feed on the tree host’s tissues directly, so they tend to develop in a more restricted number of hosts. Both can be transported in almost all kinds of wood products, where they are protected from environmental extremes and detection by inspectors. Sitzia et al. specify the usual suspects: wood packaging and plants for planting, as ideal pathways. These invasions threaten indigenous species by shifting the distribution and abundance of certain plants, altering habitats, and changing food supplies. The resulting damage to native forests induces severe alterations of the landscape and causes economic losses in tree plantations and managed forests. The latter losses are primarily in the high costs of eradication efforts – and their frequent failure.

Eucalyptus plantation in Kwa-Zulu-Natal, South Africa; photo by Kwa-Zulu-Natal Department of Transportation

Perhaps their greatest contribution is their warning about probable damage caused by invasive forest pests in tropical forests. (See an earlier blog about invasive pests in Africa.) Sitzia et al. believe that bark and ambrosia beetles introduced to tropical forests threaten to cause damage of the same magnitude as climate change and clear cutting, but there is little information about such introductions. Tropical forests are exposed to invading beetles in several ways:  

1) A long history of plant movement has occurred between tropical regions. Sitzia et al. contend that the same traits sought for commercial production contribute to risk of invasion.

2) Logging and conversion of tropical forests into plantation forestry and agriculture entails movement of potentially invasive plants to new areas. Canopies, understory plant communities, and soils are all disturbed. Seeds, insects, and pathogens can be introduced via contaminated equipment.

3) Less developed nations are often at a disadvantage in managing potential invasion. Resources may be fewer, competing priorities more compelling, or potential threats less obvious.

Sitzia et al. call for development of invasive species management strategies that are relevant to and realistic for less developed countries. These strategies must account for interactions between non-native species and other aspects of global environmental change. Professional foresters have a role here. One clear need is to set out practices for dealing with conflicts between actors driven by contrasting forestry and conservation interests. These approaches should incorporate the goals of shielding protected areas, habitat types and species from bioinvasion risk. Sitzia et al. also discuss how to address the fact that many widely used forestry trees are invasive. (See my earlier blog about pines planted in New Zealand.)

planted forest in Sardinia, Italy; photo by Torvlag via Flickr

In Europe, bark beetle invasions have damaged an estimated ~124 M m2 between 1958 and 2001. Sitzia et al. report that the introduction rate of non-native scolytins has increased sharply. As in the US, many are from Asia. They expect this trend to increase in the future, following rising global trade and climate change. Southern – Mediterranean – Europe is especially vulnerable. The region has great habitat diversity; a large number of potential host trees; and the climate is dry and warm with mild winters. The region has a legacy of widespread planting of non-native trees which are now important components of the region’s economy, history and culture. These include a significant number of tree species that are controversial because they are – or appear to be – invasive. Thus, new problems related to invasive plants are likely to emerge.

Noting that different species and invasion stages require different action, Sitzia et al. point to forest planning as an important tool. Again the discussion centers on Europe. Individual states set forest policies. Two complications are the facts that nearly half of European forests are privately owned; and stakeholders differ in their understanding of the concept of “sustainability”. Does it mean ‘sustainable yield’ of timber? Or providing multiple goods and services? Or sustaining evolution of forest ecosystems with restrictions on the use of non-native species? Resolving these issues requires engagement of all the stakeholders.

Sitzia et al. say there has recently been progress. The Council of Europe issued a voluntary Code of Conduct on Invasive Alien Trees in 2017 that provides guidelines on key pathways. A workshop in 2019 elaborated global guidelines for the sustainable use of non-native tree species, based on the Bern Convention Code of Conduct on Invasive Alien Trees. The workshop issued eight recommendations:

  • Use native trees, or non-invasive non-native trees;
  • Comply with international, national, and regional regulations concerning non-native trees;
  • Be aware of the risk of bioinvasion and consider global change trends;
  • Design and adopt tailored practices for plantation site selection and silvicultural management;
  • Promote and implement early detection and rapid response programs;
  • Design and adopt practices for invasive non-native tree control, habitat restoration, and for dealing with highly modified ecosystems;
  • Engage with stakeholders on the risks posed by invasive NIS trees, the impacts caused, and the options for management; and
  • Develop and support global networks, collaborative research, and information sharing on native and non-native trees.

SOURCE

Sitzia, T., T. Campagnaro, G. Brundu, M. Faccoli, A. Santini and B.L. Webber. 2021 Forest Ecosystems. in Barker, K. and R.A. Francis. Routledge Handbook of Biosecurity and Invasive Species. ISBN 9780367763213

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

www.fadingforests.org

Eradicating Invasive Species: You need “social license” to succeed

spread of non-native conifers in mountains of New Zealand; photos by Richard Bowman; New Zealand government website

As those of us who want to “do something” to counter bioinvasions struggle to mobilize both the  resources and the political will necessary, I rejoice that more studies are examining what factors affect “social license” [= public approva] for such programs. One such study was recently published in New Zealand — Mason et al. (full citation at the end of the blog). New Zealand enjoys a greater appreciation of the uniqueness of its biology and awareness of invasive species’ impacts than the United States. However, their findings might provide useful guidance in the US and elsewhere.

Mason et al. sought to understand motivations of, and constraints on, those local groups responsible for controlling the spread of non-native conifers into New Zealand’s remnant native ecosystems. Non-forest ecosystems across much of the country are at risk of rapidly transforming into exotic conifer forests. For these reasons, authorities are pressing for timely removal of existing seed sources, that is, mature non-native conifer trees of several species. The blog I posted earlier apparently describes effects of conifer invasions in lowland ecosystems, whereas the Programme described here is focused on high-elevation systems.

The eradication effort in the study is the National Wilding Conifer Control Programme, establishedin 2016. A large increase in funding provided during the COVID-19 lockdown made it practical to try to eradicate seed sources from large swathes of vulnerable land. The Programme coordinates control efforts across the country, working across property and land-tenure boundaries. Landowners are expected to cover 20% of the cost of removing conifers from their land. Since removing all seed sources of high-risk conifer species from the landscape is key to achieving long-term goals, success is unlikely if significant seed sources are allowed to persist.

Mason et al. combined workshops, questionnaires, and site visits to gather data on particular aspects of this Programme. They found that social resistance, rather than lack of scientific knowledge, was often the main barrier to success in managing widespread invasive species. The authors do not address whether the fact that only 30 people provided information for their study might undermine the reliability of their findings.

map of conifer wilding sites; adapted from Wilding conifers – New Zealand history and research background, a presentation by Nick Ledgard at the “Managing wilding conifers in New Zealand – present and future” workshop (2003)

The authors suggest that the main benefit of scientific information might be to increase stakeholders’ support for management interventions — rather than to guide manager’ decisions about which strategies to pursue.  To support social license, invasive species research programs might need to focus not only on cost-effective control technologies and strategies, but – perhaps especially — the benefits (both tangible and intangible) of invasive species control for society.

Mason et al. found that people were motivated to combat conifer invasions by impacts with direct influence on humans or human activities (e.g., reduced water yield, damage to infrastructure from wildfires, reduced tourist activities due to landscape transformation) and also by impacts affect ecosystems (e.g.,  impacts on biodiversity, aquatic ecosystems and landscapes).  

People objected to control or eradication programs primarily because of social concerns. These included the unwillingness of landowners to participate and regulatory frameworks that had perverse incentives.

Mason et al. called for greater efforts by scientists to persuade stakeholders[p1]  to allow removal of “wilding” conifers from private land and development of more appropriate regulations. They found that forecasting models were particularly effective in persuading people to support these efforts. It seems to me that outreach teams might need “translators” to convert scientists’ findings to information that would be more useful by stakeholders.

The authors concede that the “wilding conifer” situation has unique attributes. First, invading conifers present a stark, easily seen difference between native and invaded ecosystems. Second, some – but not all—stakeholders appreciate the uniqueness of New Zealand’s biomes. Third, the impacts of conifer invasion are sufficiently well known that they can be described succinctly and accurately.

Do these unique attributes undercut the relevance of this study to North America? It is still true that ongoing support from local stakeholders (including landowners and community groups) influences the effectiveness or profitability of managing invasive species. .It is also true that groups’ varying values affect willingness to support the activities.

Mason et al. think through the issue of stakeholders’ conflicting perspectives on the value of particular invasive species and the values threatened by that invader. These can include ethical or safety concerns around management methods, particularly regarding toxins and genetic modification. Economoic costs are also a factor – especially if the landowner must pay all or some of them.

I find it interesting that the government simultaneously funded a 5-year research program to study various issues regarding the spread, ecosystem impacts, and control of wilding conifers. The result is the Mason et al. study discussed here. I wish the U.S. funded independent analyses of its invasive species programs!

Pinus contorta – the most rapidly growing Pinus introduced to New Zealand; photo by Walter Siegmund / Wikimedia

More Details, Policy Suggestions

Workshop attendees unanimously identified landscape impacts as a reason for controlling wilding conifers. This primarily concerned the loss of New Zealand’s visual heritage or cultural identity rather than loss of native species’ habitats. When the landowner was raised in Europe, these cultural or heritage values sometimes had the opposite effect, since they see conifer forests as important components of “natural” landscapes.

Currently, landowner funding and permission is required for conifer removal. Some individual landowners want to establish new forestry plantings. Some resist removal of existing forestry plantations (which provide income) and shelter belts (which provide shelter for livestock in high country landscapes). Some landowners were unwilling to pay their 20% of removal costs. Or they objected to certain conifer control methods—particularly helicopter spraying of herbicides. New Zealand’s regulatory process also requires years of negotiating to remove standing trees – further delaying any action. In theory, landowners who resist removal could be prosecuted under the Biosecurity Act. However, this approach has never been tried for removing wilding conifers.

Mason et al. suggested several changes in policy to overcome some of these barriers.

First, forestry consultants can “game” the wilding conifer “risk calculator” to obtain government approval to establish conifer plantations in high-risk environments. The authors suggest that authorities create a “liability calculator.” Under this system, landowners wishing to retain conifers on their land for whatever reason would be liable for any subsequent containment costs. However, developing such a tool requires more finely-scaled models of conifer spread.

Second, given the high costs of combatting invading conifers if seed sources are allowed to persist, they suggested that it might be more cost-effective for the control program to pay for plantation removal under New Zealand’s Emissions Trading Scheme.

Given the overwhelmingly social and regulatory nature of barriers to success, the primary role for scientific information is providing assessments of outcomes in the absence of wilding conifer control. Preferred messages were return-on-investment estimates and forecasts of ecosystem impacts, particularly relating to biodiversity loss, water yield reduction, and wildfire hazard. Forecasts were key to demonstrating that management interventions reduced future control costs and avoided environmental impacts which large sections of the community value (i.e. biodiversity loss, reduction in water yield and agricultural productivity, increased wildfire risks). Practitioners felt that forecasting models might also channel research toward areas of high uncertainty. Mason et al. recognize the difficulties presented by inherent complexity of ecological systems. However, they think “good practice” guidelines on forecasting are emerging.

The authors find that information content and presentation need to be tailored to the various audiences – most of whom lack experience in interpreting data from environmental forecasting models. They suggest that outreach materials focus on clear illustration of the tangible and intangible benefits of wilding conifer management rather than detailed explorations of scenarios. Participants suggested ways to improve the web tool to make it more accessible to a non-expert audience. 

Mason et al. mention aspects that require balancing, but don’t suggest criteria for making these choices. They say it is important to include all relevant stakeholders in invasive species management governance bodies. The absence of stakeholders with positive attitudes to wilding conifer invasions led to unanticipated external social resistance to the Programme. They recognize that including stakeholders with conflicting interests might obstruct the decision-making process.  Also, in areas where there has been success in containing conifers’ spread, people can’t see invading trees, so they don’t recognize the problem. They also note that existing data do not adequately recognize risks of spread from deliberately planted seed sources such as shelter-belts, plantations and amenity plantings. The authors do not discuss how to integrate these data into analyses and public outreach.

Finally, Mason et al. recognize that many other factors strongly influence stakeholders’ willingness to support invasive species control programs, especially the level of trust and strength of relationships between bioinvasion program staff and stakeholders.

Also, they suggest topics for future research: assessing how well forecasting models are integrated with  communications with stakeholders; how qualitative and quantitative research methods in different fields might support one another; and empirical tests to measure the relative effects on social license of a) involving stakeholders in developing models, b) using forecasts to assess the consequences of different management decisions and, c) the usefulness of different methods for incorporating scientific information in stakeholder engagement.

SOURCE

Mason, N.W.H., Kirk, N.A., Price, R.J. et al. Science for social license to arrest an ecosystem-transforming invasion. Biol Invasions 25, 873–888 (2023). https://doi.org/10.1007/s10530-022-02953-w

see also https://www.doc.govt.nz/nature/pests-and-threats/weeds/common-weeds/wilding-conifers/

Posted by Faith Campbell

What do YOU think about the role “social license” plays in US invasive species programs? 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

www.fadingforests.org

 

Climate Change + CO2 Levels – Can Scientists Include the Complexity in their Analyses?

Spruce budworm (Choristoneura fumiferana); photo by Jerald E. Dewey, USFS; via Bugwood; populations of several forest birds, including Cape May, Tennessee and Bay-Breasted warblers, become more numerous during budworm outbreaks

Now that Drs. Ziska and Aucott have educated us about the strong impact atmospheric CO2 can have on both plants and phytopagous insects, I have asked the experts whether these interactions have been incorporated in the models scientists are using to forecast pest activity in American forests as the climate changes.

The answer is no.

bay-breasted warbler; photograph by Dave Inman at Presque Isle State Park, PA; via Flickr

Dr. Bethany A. Bradley, Co-Director, Northeast Climate Adaptation Science Center at the University of Massachusetts, says empirical models of species range shifts typically only use climate and sometimes other environmental factors (like soils or topography) as predictors of potential geography. Inclusion of demographic processes like how plant growth is affected by more or less water, CO2, competition with other plants etc. would require a lot of data. It is currently impossible since there are tens of thousands of plant species interacting in the forests of eastern North America – and perhaps these factors have been analysed for only a hundred of them.

Mike Aucott points to the same difficulty: inclusion of CO2 in models of the future populations of specific plants would be difficult since the impacts vary from species to species and are compounded by other factors such as soil nitrogen levels, moisture levels, temperature, presence of competing plants, etc.  

Regarding insects, Dr. Aucott thinks it is clear that some orders, such as Lepidoptera, don’t fare as well when feeding on plants grown under elevated CO2.  He is not aware of efforts to model impacts of high CO2 on specific insects or even orders or feeding guilds. 

juniper geometer (inchworm); Dr. Tallamy says inchworms are hairless & good tasting – so sought by birds

Dr. Ziska concurs about the difficulties. Dr. Ziska asks why there is so little funding to study these issues, especially given their probable impact on human food supplies and health – as described in his blog and an opinion piece published in Scientific American two years ago.

I hope that scientists, decision-makers, readers of this blog … maybe even the media! – take into consideration these complexities, even if they cannot be defined.

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 – [but do not address climate or CO2 aspects] review the Fading Forests report at http://treeimprovement.utk.edu/FadingForests.htm

or

www.fadingforests.org

Global Weirding: Rising CO2 Impacts Plants & People

Guest blog by Lewis Ziska, Associate Professor, Environmental Health Sciences at the Columbia University

[Dr. Ziska has spent his career analyzing the impacts of CO2 and climate change on plants – and therefore on people. He served as Project Leader for global climate change at the International Rice Research Institute; then spent 24 years at the USDA’s Agricultural Research Service, where he worked primarily on documenting the impact of climate change and rising carbon dioxide levels on: Crop selection improves production; Climate and agronomic pests, including chemical management; Climate, plant biology and public health impacts on food security with a focus on nutrition and pesticide use.]

No question you’ve heard the term, “Climate Change” or “Global Warming”, or my personal favorite, “Global Weirding”. The consequences are talked and discussed in the media—as they should be—but often the media, like many Americans, is focus challenged. Or in more polite terms, they have the attention span of a hummingbird on crack. Which is to say, that simple physical consequences, like sea level rise (heat melts ice!), and stranded animals on ice (Poor polar bear!), or intense storms (newscaster whipped about in the rain, yelling to be understood) are repeated, over and over again. Understandable, makes for good TV.

But it also makes you feel separate from what is happening, these consequences of climate change are to the “other”. I don’t live near the ocean, I don’t interact with polar bears; sure we have storms, but I live in the Midwest, in one of those states that begins with a vowel. Shoot, I commute to work, try and make ends meet, I’m not some damn tree hugger. Why should I care? 

To understand why, you need a bit more background, some science that isn’t always available on TV or social media when it comes to global weirding.

First, while you may not be a tree hugger, you do, in fact, interact with nature. Several times. Every day. We call those times, “breakfast”, “lunch” and “dinner”.

dinner; photo by davylin via Flickr

You depend on nature for food. And clothing. And paper. And medicine. And oxygen. And construction materials (wood), and many, many other things. So, if nature gets hinky, and the climate becomes uncertain, it might be worth your while to think about climate change, or global weirding, in a different light. What I want to do here then, is to illuminate two examples that I hope will help you see why climate could affect you, directly and significantly.

Let’s begin with plants. Those green living things that comprise the bulk of the natural world (literally, if you were to weigh the natural world, 97% would be plants, 3% animals). Then let’s look at them through two different lenses—how will climate weirding alter your food; shoot, how will it alter the air that I breathe?

Let’s start with a basic food, rice. Obviously you don’t want to mess with its production, or its nutritional quality. But that is exactly what global weirding is doing.

Rice has flowers. Not big showy ones, but flowers none the less—ones that get fertilized with pollen, and seed is produced. The seed that feeds some two billion people– or about a quarter of the earth’s population.

Like all living things, plants are heat sensitive, and for rice, and many crop plants, the degree of sensitivity varies, depending on the part of the plant in question. Take a look at the table. The crops that are listed, including rice, are the core of what the world eats. Now notice the difference in temperature sensitivity. Vegetative parts of the plant, leaves and stems, are reasonably tolerant of higher temperatures, but flowers are not. Pollen, the plant equivalent of animal sperm, is highly temperature sensitive, and if the temperatures get into the high 90s (37-38oC), they become deformed, and the rice plant doesn’t produce seeds. Same for a number of plants, ones necessary to feed 8 billion people.

CropOpt. Temp. VegetativeOpt. Temp. FloweringFailure Temp. Flowering
Maize 28-35oC 18-22oC 35oC
Soybean 25-37oC 22-24oC 39oC
Wheat 20-30oC  15oC 34oC
Rice 28-35oC 23-26oC 36oC
Sorghum 26-34oC  25oC 35oC
Cotton  34oC 25-26oC 35oC
Peanut 31-35oC 20-26oC 39oC

Data are adapted from Hatfield et al., 2011.

Doubtful you’ve seen this climate threat to the global food supply on TV or a streaming service. I caught a glimpse once of temperature and agriculture on a CNN newscast, but with the “expert” calmly stating that we would just have to grow our corn in Canada, ha-ha. (Somehow, at least for rice, it’s hard to imagine India, one of the world’s largest rice producers, moving its rice production northward to the Himalaya’s, but I digress.)   

Food is fundamental. If production, especially that of a global staple like rice, is impacted by rising temperatures there will be consequences. Rising prices, reduced availability, and wide-spread hunger.

But there is more to consider. Given the global dependence on rice, any change in its nutritional quality will also have effects, especially on poorer countries that rely heavily on rice as a major food source. And here we need to delve a little deeper into another aspect of climate weirding that doesn’t make it to the popular media—that rising carbon dioxide (CO2), the primary greenhouse gas, can also directly influence plant nutrition.  The reasons are complicated, but in simple terms all living things consist of elements, carbon, nitrogen, phosphorous, sulfur, copper, etc., etc. A plant gets it’s carbon from the air (CO2), but everything else (nitrogen, potassium) from the soil.

And there is an imbalance. In the last 50 years, atmospheric CO2 has increased by about 30%, and is projected to increase another 50% by the end of the century. With more CO2, plants are becoming carbon rich, but nutrient poor. Nutrient poor, because while CO2 has increased in the air, nutrients in the soil have not kept pace. A perverse carb loading at the plant level.

As a consequence, rice, and many other plants, are shifting their chemistry. For example, there is a general decline in protein, in part because protein requires nitrogen. There are similar ubiquitous declines in iron and zinc, important micro-nutrients needed for human development.

Such nutritional degradation is of obvious global importance, and does, on occasion, show up on basic media when warming / weirding is mentioned, but you’d be hard pressed to find it.

Let’s move our light to another hidden bit of science. How plants can influence the air we breathe.

As humans, we like to trade things. And a large percentage of what we trade are living organisms, from fish to trees.  But what began as local, regionalized trading has grown with the global population and the needs of that population—a population of 1.6 billion at the beginning of the 20th century is now ~8 billion at the beginning of the 21st. And we haven’t stopped trading. Biological trade is not inherently bad, but it represents a historically unprecedented global movement of DNA across continents, across countries, regions, towns, cities and ecosystems. And some of the DNA, when introduced, can do great harm to the environment, the economy and to human health. That harm has a name, “Invasive Species”.

Let us focus on one such plant species introduced to Eastern Europe, one that almost every American has personal experience (ACHOO!) come fall. The species is common ragweed. An invasive plant whose introduction and spread in Eastern Europe—introduced accidently through imported seeds or contaminated hay – has resulted in enormous environmental and economic losses in agriculture and public health in recent decades. In Hungary, the most important ambient biological air pollutant is: ragweed.

collecting ragweed pollen under different climates (Author’s photo)

The photo is from studies that I led looking at how ragweed pollen would respond to temperature and carbon dioxide. (If you’re curious, ragweed likes both.)  Warmer temperatures, earlier Springs, later Autumns can extend its pollen season; not only extend, but increase the amount of pollen being generated. There is even some data suggesting that rising CO2 can alter pollen chemistry, making it more allergenic (REFS). Sadly, ragweed pollen doesn’t appear as temperature sensitive as that of rice, or other agricultural plants.

I wish I could say that ragweed was the exception among allergenic plants, but it’s the rule. Parthenium weed is a highly invasive species that has spread to more than 40 countries around the world. Like ragweed its pollen are highly allergenic, but it can also produce severe rashes, like poison ivy, and is known to be poisonous to livestock. It is highly aggressive, and arriving in a new location (where it has no natural enemies) can dominate landscapes, reducing biodiversity. And as with ragweed, high temperatures, longer growing seasons, heatwaves and droughts are expanding its range, and for that matter, make controlling its spread more difficult.

Such responses among invasive species will have direct impacts on air quality, especially among those (myself included) who suffer from seasonal allergies. Gasping for air is never fun.

Estimates are that pollen and seasonal asthma affects more than 24 million of us, including 6 million kids. And yet, when watching news reports of climate change, how many times have you seen a report on pollen and air quality?  On increasing allergies or asthma?  Once?  Twice? 

I could go on, (and if you need more examples, read “Greenhouse Planet”, my latest book). But my point is this:  Not all of the consequences of rising carbon dioxide and climate change, warming, weirding, whatever, make for “good” TV. There is so much more to explore. So, do yourself a favor. Take a deeper dive, find out what is happening behind the scenes.

Because if we are going to rise to the challenge, we need to know what we are fighting against. Right now, the media is exemplary on showing some things, but silent on much else of importance. Watching news coverage of climate change is a bystander watching a cataclysm, and thinking, “Boy, glad I’m not experiencing THAT!”. Yet in the simplest and most basic of terms, you are, or will be, affected– from food choices to nutrition, even your allergies. And so much more.

It isn’t just about polar bears. It’s about you. Read, Understand, Act.

Now.  

EAB in Eastern Europe – Worse (+ war!)

A special issue of the journal Forests (Vol. 13 2022) seeks to improve understanding of the root causes of exacerbated threats from insect pests. The issue contains 15 papers; most focus on geographic areas other than North America. The journal is open access!

Choi and Park (full citations below) link increased pest risk to climate change and increased international trade. They provide brief summaries of all 15 papers. My focus here is on two articles that provide updates on the status of the emerald ash borer (EAB Agrilus planipennis) in Russia and Ukraine. The article by Davydenko et al. also examines interactions between EAB and the invasive pathogen Hymenoscyphus fraxineus, which causes ash dieback disease. In other blogs I will look at insects linked to North America (both species from North America that threaten forests on other continents, and species in Russia that pose a threat to North America) and at the overall Russian experience.

I blogged about EAB invasion of Russia in April 2021 so this is an update.

Musolin et al. (2022) (full citations below) remind us that the EAB invasions of North America and Russia were detected almost simultaneously: in Michigan and Ontario in 2002 and in European Russia (Moscow) in 2003. They conclude that both invasions probably originated from a common source (most probably China). They date the introduction to the late 1980s or early 1990s; pathways might have been wooden crafts, wood packaging, or ash seedlings. Nate Siegert used dendrological studies to estimate a similar introduction date for the North American invasion.

European ash (Fraxinus excelsior) specimen in Belgium; photo by Jean-Pol Granmont

EAB has spread far in the intervening 30 + years. By early 2022, outbreaks were recorded in five Canadian provinces, 35 US states, 18 provinces and several cities in European Russia, and two provinces in Ukraine (Musolin et al. 2022) Davydenko et al. report that EAB had also established in eastern Belarus, but provide no details.

As demonstrated in the earlier blog and confirmed by Musolin et al. (2022) and Davydenko et al., the EAB has spread much faster to the southwest than directly West and to the Northwest. Davydenko et al. attribute the slower spread in the St. Petersburg area to the colder and wetter climate of this region – which is ~1200 km north of Ukraine. While the EAB reproduces in two cohorts in Eastern Ukraine, to the north the beetle requires more than one year to complete its life cycle, at least two years in the St. Petersburg area. In 2021, Musolin et al. 2021 speculated that pressure by the native parasitoid Spathius polonicus Niezabitowski might also be slowing EAB’s spread in the North. In 2022, Musolin does not address this possibility. (I note that APHIS has approved two Spathius species as biocontrol agents in the U.S.).

Musolin et al. (2022) and Davydenko et al. agree that the EAB poses real threat to ash in central and western Europe. In both the south (Davydenko et al.) and in the northwestern area around St. Petersburg ash grows in continuous stretches, linking Russia or Ukraine to Romania, Hungary, Slovakia, and Poland. These ash consist of both natural woodlands, and extensive plantings of both one of the European ash species, F. excelsior and the highly-susceptible North America green ash (F. pennsyvanica).  Furthermore, the EAB is an excellent hitchhiker on vehicles & railway cars. Davydenko et al. also consider the beetle to be a strong flyer. Musolin et al. (2022) cite a separate analysis in stating that EAB can probably invade most European countries. Only some regions of Norway, Sweden, Finland, Ireland, and Great Britain are probably protected by their low temperatures.

Both articles were written too early to consider how the current war in the relevant area of Ukraine might affect spread of the EAB, although we know Ukrainians are cutting firewood. The war has certainly interrupted monitoring and other efforts.

The sources agree on EAB’s severe impacts. Musolin et al. (2022) notes that the beetle has killed millions of trees in the forests and urban plantings in North America, European Russia, and Eastern Ukraine. Davydenko et al. note that the Fraxinus genus is one of the most widely distributed tree genera in North America. They then assert that the EAB could virtually eliminate it. I know that North American scientists agree that the beetle threatens many species in the genus; but do they agree that the genus would be “virtually eliminated”?  Davydenko et al. think the EAB could pose similar threat to Euro ash F. excelsior.

Musolin et al.  2022 estimate that potential economic losses in Europe could reach US$1.81 billion. By this indicator, the species ranks fourth among the most “costly” invasive pests. Russia spent an estimated US$258.9 million between 2011 and 2016.

areas of Ukraine where studies conducted

Species’ varying vulnerability

Musolin et al. (2022) cite experience in the Moscow Botanical Garden as showing that only two Asian species — Chinese ash, F. chinensis, and Manchurian ash, F. mandshurica — are were resistant to the EAB. The beetle killed both North American ash (i.e., F. pennsylvanica and F. americana) and European ash (i.e., F. excelsior, F. angustifolia, and F. ornus).

Experience in the field in Ukraine (Davydenko et al.) suggests that F. excelsior is less vulnerable to EAB than F. pennsyvanica. The overwhelming majority of EAB infestations were found on the American species. Furthermore, although similar densities of EAB larvae were found in colonized branches of both species, the proportion of larvae that were viable was significantly higher on F. pennsyvnica (91.4%) than on F. excelsior (76.1%). However, the reverse was found in the Moscow and St. Petersburg regions. Davydenko et al. don’t address directly whether they think this discrepancy is attributable to climatic factors or to differences in vulnerability between trees growing in native forests vs. human plantings. They did note that all observed cases of infestation of the native F. excelsior in Ukraine occurred in artificial plantings rather than in natural woodlands.  

Interactions with Ash Decline

ash dieback disease – unfortunately pictured in the UK.
cc-by-sa/2.0 – © Adrian Diack – geograph.org.uk/p/6497286

Davydenko et al. studied parts of Eastern Ukraine where EAB was entering areas already infected by the invasive ascomycete fungus Hymenoscyphus fraxineus (cause of ash dieback, ADB). [Two of these regions — Luhansk and Kharkiv – have been the very center of the current war.] Other studies have shown that ~1 to 5% of F. excelsior trees exhibit some resistance to ADB. These trees are thus a potential foundation for future propagation and restoration of ash in Europe – if enough of them survive attack by EAB.

They found that F. excelsior is more resistant to EAB than F. pennsylvanica, but more susceptible to ADB.

The Luhansk and Kharkiv regions have both EAB and ADB; the Sumy region has only the pathogen. EAB probably invaded the Luhansk region by 2016 (although it was detected only in 2019). The proportion of ash trees (both native and introduced species) infested rose from ~ 10–30% in 2019 to 60 – 90% by 2020–2021. The EAB arrived later in the Kharkiv region, to the Northwest, but the proportion of infested trees was similar by 2021. Combining the two regions, 75% of F. pennsylvanica trees were EAB-infested, whereas only 31% of F. excelsior trees were.

Frequencies of infections by ADB were the reverse. Pooled data from all three study regions showed 21% of F. pennsylvanica trees were infected vs. 42% of F. excelsior. In the plots invaded by both EAB and ADB (in Luhansk and Kherson regions), 4%of F. pennsylvanica were affected by both invasive species vs. 14% of F. excelsior trees. Davydenko et al. conclude that ADB facilitates EAB attack on F. excelsior trees

The impact of EAB is seen in the fact that overall mortality rates were higher in F. pennsylvanica despite the fact that in the Sumy region mortality rates were higher in F. excelsior because of the disease (EAB was absent from this region).  On the other hand, EAB infests and kills F. pennsylvanica trees regardless of their prior health condition (i.e., regardless of presence/absence of ADB).

Still, fewer than half the F. excelsior trees in sites affected by both EAB & ADB (in Luhansk and Kherson regions) have died. Davydenko et al. think the survivors constitute a source of material for eventual propagation. These trees need to be carefully mapped – a task certainly not facilitated by the war!

Davydenko et al. conclude that

1. Invasion of EAB in Ukraine occurred 2–3 years before detection in 2019 [I think this is actually quite prompt for detection of EAB invasions]; the invasion is currently expanding both in terms of newly infested trees and invaded geographic area.

2. Fraxinus excelsior (at least when growing in the interior of forest stands) is more resistant to EAB than F. pennsylvanica (when growing in field shelterbelts).

3. Fraxinus excelsior is more susceptible to ADB than F. pennsylvanica.

4. Infection by ADB is likely to predispose F. excelsior to infestation by EAB.

5. Ash trees infected by ADB are predisposed for the colonization by ash bark beetles Hylesinus spp.  [I did not discuss these data.]

6. Inventory and mapping of surviving F. excelsior, affected by both ADB and EAB, is necessary to acquire genetic resources for the work on strategic, long-term restoration of devastated areas, thereby tackling a possible invasion of EAB to the EU.

I was surprised that Musolin et al. (2022) think EAB’s host shift from local Asian ash species to introduced North America ash planted in the Russian Far East and China triggered EAB outbreaks in Eastern China that contributed to the beetle’s introduction to North America and European Russia. American scientists apparently agree — Haack et al. (2022) refer to both this episode and a similar to one posited for Asian longhorned beetle (Anoplophora glabripennis) — that widespread planting of Populus plantations led to rapid expansion of ALB in northern China, and the pest-weakened wood was then used in wood packaging.

SOURCES

Choi, W.I.; Park, Y.-S. Management of Forest Pests and Diseases. Forests 2022, 13, 1765. https://doi.org/10.3390/f13111765

Davydenko, K.; Skrylnyk, Y.; Borysenko, O.; Menkis, A.; Vysotska, N.; Meshkova, V.; Olson, Å.; Elfstrand, M.; Vasaitis, R. Invasion of emerald ash borer Agrilus planipennis and ash dieback pathogen Hymenoscyphus fraxineus in Ukraine-A concerted action. Forests 2022, 13, 789.

Haack RA, Hardin JA, Caton BP and Petrice TR (2022) Wood borer detection rates on wood packaging materials entering the United States during different phases of ISPM#15 implementation and regulatory changes. Front. For. Glob. Change 5:1069117. doi: 10.3389/ffgc.2022.1069117

Musolin, D.L.; Selikhovkin, A.V.; Peregudova, E.Y.; Popovichev, B.G.; Mandelshtam, M.Y.; Baranchikov, Y.N.; Vasaitis, R. North-Westward Expansion of the Invasive Range of EAB, Agrilus planipennis Fairmaire (Coleoptera: Buprestidae) towards the EU: From Moscow to Saint Petersburg. Forests 2021, 12, 502. https://doi.org/10.3390/f12040502

Musolin, D.L.; Kirichenko, N.I.; Karpun, N.N.; Aksenenko, E.V.; Golub, V.B.; Kerchev, I.A.; Mandelshtam, M.Y.; Vasaitis, R.; Volkovitsh, M.G.; Zhuravleva, E.N.; et al. Invasive insect pests of forests and urban trees in Russia: Origin, pathways, damage, and management. Forests 2022, 13, 521.

Siegert, N.W.  2006.  17th USDA Interagency Research Forum on Gypsy Moth and Other Invasive Species. Annapolis, MD. January 10-13, 2006.

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

www.fadingforests.org

US invasive species — updated USGS database now on-line

ōhiʻa rust on Hawai`i; photo by J.B. Friday

The U.S. Geological Survey (USGS) has published an updated register of introduced species in the United States. The master list contains 14,700 records, of which 12,571 are unique scientific names. The database is divided into three sub-lists: Alaska, with 545 records; Hawai`i, with 5,628 records; and conterminous (lower 48) United States, with 8,527 records.

The project tracks all introduced (non-native) species that become established, because they might eventually become invasive. The list includes all taxa that are non-native everywhere in the locality (Alaska, Hawai`i, or 48 conterminous states) and established (reproducing) anywhere in that locality.

Each record has information on taxonomy, a vernacular name, establishment means (e.g.,  unintentionally, or assisted colonization), degree of establishment (established, invasive, or widespread invasive), hybrid status, pathway of introduction (if known), habitat (if known), whether a biocontrol species, dates of introduction (if known; currently 47% of the records), associated taxa (where applicable), native and introduced distributions (when known), and citations for the authoritative source(s) from which this information is drawn. 

The 2022 version is more complete re: plant pathogens than earlier iterations; I thank the hard-working compilers for their efforts!

Hawai`i

wiliwili tree (Erythrina sandwicensis); photo by Forest and Kim Starr

Among the non-native species listed as being in Hawai`i are 3,603 Arthropods, including the following about which I have blogged:

The list also includes 25 fungi, among them the two species of Ceratocystis that cause rapid ʻōhiʻa death; DMF & blog 270 and the ʻōhiʻa or myrtle rust, Austropuccinia psidii.

Also listed are 95 mollusk species and 20 earthworm species. I wonder who is studying the worms’ impacts? I doubt any is native to the Islands.

The Hawaiian list contains 1,557 non-native plant species. Families with largest representation are Poaceae (grass) – 223 species; Fabaceae (beans) – 156 species; and Asteraceae – 116 species. About a third of the plant species – 529 species – are designated as “widespread invaders”. This number is fifteen times higher than the numbers in lists maintained by either the Hawaiian Ecosystems At Risk project (106 species) [HEAR unfortunately had to shut down a decade ago due to lack of funds]; or Hawaiian Invasive Species Council (80 species). Furthermore, some of the species listed by HEAR and HISC are not yet widespread; the lists are intended to facilitate rapid responses to new detections.  We always knew Hawai`i was being overrun by invasive species!

Among the 529 most “widespread invaders” are the following from the most introduced families:

  • Poaceae – Agrostis stolonifera, 6 Cenchrus spp, 2 Cortaderia spp, 3 Eragrostis,8 Paspalum, 4 Setaria spp, 2 Urochloa (Poacae)
  • Fabaceae – 3 Acacia, 2 Prosopis

Other families have fewer introduced species overall, but notable numbers of the most widespread invaders:

  • Euphorbiaceae – 8 spp. of Euphorbia
  • Cyperaceae – 6 spp. of Cyperus
  • Myrtaceae – Melaleuca quinquenervia, 2 Psidium, Rhodomyrtus tomentosa rose myrtle, 3 Syzygium [rose myrtle has been hard-hit by the introduced myrtle rust fungus]
  • Zingiberaceae – 3spp. Hedychium (ginger)
  • Anacardiaceae — Schinus molle (Peruvian peppertree); USGS considers congeneric S. terebinthifolia to be somewhat less widespread.

Plus many plant taxa familiar to those of us on the continent: English ivy, privet, castor bean, butterfly bush, Ipomoea vines  … and in more limited regions, Japanese climbing fern Lygodium japonicum.

Rhus sandwicensis; photo by Forest and Kim Starr

I learned something alarming from the species profiles posted on the HISC website: the Hawaiʻi Division of Forestry and Wildlife and Hawaiʻi Department of Agriculture are considering introduction of a species of thrips, Pseudophilothrips ichini, as a biocontrol agent targetting S. terebinthifolia. I learned in early 2019, when preparing comments on Florida’s proposed release of this thrips, that Pseudophilothrips ichini can reproduce in low numbers on several non-target plant species, including two native Hawaiian plants that play important roles in revegetating disturbed areas. These are Hawaiian sumac Rhus sandwicensis and Dodonea viscosa. The latter in particular is being propagated and outplanted in large numbers to restore upland and dryland native ecosystems. While the environmental assessment prepared by the USDA Animal and Plant Service says the thrips causes minimal damage to D. viscosa, I am concerned because of the plant species’ ecological importance.  Of course, the two Schinus species are very damaging invasive species in Hawai`i … but I think introducing this thrips is too risky. [To obtain a copy of CISP’s comments, put a request in comments section. Be sure to include your email address in your comment; the section algorithm does not include email addresses (how inconvenient!).]

Continental (lower 48) states

Among the 8,500 species listed in the USGS Register for the 48 continental states are 4,369 animals, among them 3,800 arthropods; 3,999 plants; and just 89 fungi. Among the arthropods, there are 1,045 beetles and 308 lepidopterans. The beetles listed include 12 Agrilus (the genus which includes emerald ash borer and goldspotted oak borer.) It does not include the elm zig-zag sawfly USGS staff have not found any publications documenting its U.S. occurrences. Among the microbes are six Phytophthora (P. cinnamomi, P. lateralis, P. pseudocryptogea, P. quercina, P. ramorum, P. tentaculata). Profiles of several of these species are posted at www.dontmovefirewood.org; click on “invasive species”, then scroll using either Latin or common name.

elm zig-zag sawfly; photo by Gyorgy Czoka via Bugwood

Citation:

Simpson, Annie, Pam Fuller, Kevin Faccenda, Neal Evenhuis, Janis Matsunaga, and Matt Bowser, 2022, United States Register of Introduced and Invasive Species (US-RIIS) (ver. 2.0, November 2022): U.S. Geological Survey data release, https://doi.org/10.5066/P9KFFTOD

United States Register of Introduced and Invasive Species; US-RIIS ver. 2.0, 2022

 If you would like to contribute to future versions of the US-RIIS, please email the project leaders at us-riis@usgs.gov. 

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

www.fadingforests.org

Plants Sold – Increasing % of Exotics

carnation (Dianthus caryophyllus) photo by Noordzee23 via Wikimedia; carnations have been big sellers for 250 years

Plants sold in nurseries directly influence urban landscapes by providing gardens and other habitats that support humans and birds, insects, and other organisms. Doug Tallamy, though, has described ways that non-native plants fall short in providing habitat for native wildlife. Of course, non-native plants also indirectly influence natural landscapes by acting as a major source of invasive species. [see blog – includes links to regional invasive plant lists; and here] Imported plants also can carry non-native insects and pathogens – about which I blog frequently! To review these blogs, scroll down below the archives to the “categories” section and click on “plants as pest vectors”.

Now Kinlock, Adams, and van Kleunen (full citation at the end of this blog) have published a new paper that sheds more light on these issues. They analyzed the ornamental plants sold in US nurseries over 225 years (from 1719 to 1946). Their database, drawn from an earlier publication by Adams (see Sources at end of blog), included records of 5,098 ornamental vascular plant species offered for sale by 319 US nursery catalogs published over this period.

They note that present-day urban yards in the continental United States are planted in a diverse array of plants and the plants are predominately non-native species. Also, there is relatively little variation in species planted from one region to another, especially when compared to regional variation in natural areas). These patterns reflect the history of US horticulture.

Seventy percent (3,587) of the 5,098 ornamental vascular plant species offered by the 319 nurseries over those 200 years were non-native to the continental United States. They believe that the number of non-native species offered for sale has probably continued to increase in the 70 years since their study ended. They cite a study showing that 91% of tree species sold by nurseries in southern California during the 20th and early 21st centuries were not native to that state. A similar figure comes from a study of cultivated plants grown in Minneapolis–Saint Paul. There 66% of plants were non-native. (Kinlock, Adams, and van Kleunen note that 70% of species cultivated in yards of five British cities are non-native. In contrast, only 23% of cultivated plants in 18 Chinese cities were non-native.)

Kinlock, Adams, and van Kleunen note that two examples of non-native plants that have become invasive were among most common species available from nurseries beginning in the mid-19th Century: Japanese honeysuckle (Lonicera japonica) was available in 78 nurseries, and Japanese barberry (Berberis thunbergia) in 46 nurseries.

Japanese honeysuckle; photo by Chuck Bargeron, Bugwood

 Historical Trends

The earliest commercial horticulture in colonies that became the United States was in the mid-17th Century. It involved imports of Eurasian fruit trees to establish orchards to provide familiar foods. Ornamental horticulture became popular earlier than I expected. Prince Nurseries was established in 1732 in Flushing, NY. It was followed by additional nurseries in New York, Philadelphia, and Massachusetts. Originally these businesses imported Old World nursery stock and seeds – again to provide familiar foods and take advantage of relationships with European contacts.

Nurseries proliferated in the 1820s in the population centers of the Atlantic coast. As people of European ancestry moved west, so did nurseries. Kinlock, Adams, and van Kleunen point out an interesting aspect of these changes: proliferation of both was aided by technology: steamboats, canals, highways, and improved mail service. Before 1800, nearly all nurseries were in the Mid-Atlantic, New England, and South. Nurseries appeared in the Great Lakes region by the 1830s. Expansion of rail lines connected nurseries from coast to coast by the 1870s. By 1890, there were more than 4,500 nurseries across the continent.

California, Florida, and Oregon are now the states with the most horticultural operations and sales (as of 2019).

The types of plants offered for sale proliferated throughout the 19th Century.  The species richness of US nursery flora peaked in the early 20th Century. It decreased in the 1925 – 1946 period, possibly attributable to some combination of war-related interruptions to trade and a shift in gardeners’ focus away from ornamentals to vegetables. Another factor was adoption of international and interstate phytosanitary regulations in the early 20th Century. The post-World War II economic boom led to a new diversification of US nursery flora. In one study, a Los Angeles nursery experienced the largest increase in species richness during 1990–2011. They believe this increase was probably matched across the country. Global plant collection and importation mediated by US botanical gardens and nurseries remain active.

planting of Eucalyptus seedlings in California during 1980s; National Archives photo

Over time, nursery floras in the various regions became more similar to each other. The floras of Mid-Atlantic and New England nurseries differed before 1775, then became similar. Nurseries in the Great Lakes region also shifted toward offerings in neighboring regions. Later, nurseries in the South and West also began offering a higher proportion of species commonly sold across the continent. The nursery floras of Great Lakes and Great Plains regions were consistently similar. Still, the flora in Western nurseries still retain some unique aspects. California is the only state with a Mediterranean climate. Nurseries there sought adapted plant species, especially from an entirely new source — Australasia. (The authors note that Acacia and Eucalyptus genera, while important in California horticulture, are invaders in Mediterranean zones worldwide.) One might expect the need for plants in the Southwest to be drought-tolerant would prompt a unique nursery flora. However, the ubiquity of irrigation since the late 19th Century has blunted this necessity. Still, nursery flora in the desert biome had the most phylogenetic uniformity. The article does not discuss pressure to choose xeriscapes or otherwise adjust to current water shortages.

Pinus mogu – sought for xerescapes; photo by Krysztof Ziamk Kenraiz via Wikimedia

Growing Importance of Non-Native Species – Especially from Asia

Kinlock, Adams, and van Kleunen define “native” species as those native to the state in which it is sold; “adventive” species as native to the continental United States but not the specific state; and non-native or alien species as not native to the continental United States.

Applying these definitions, the proportion of native species in nursery flora has been consistently around 30-40% — except during the American Revolution. It rose to 70% in catalogs or advertisements published from 1775 to 1799. The authors do not speculate whether this reflected jingoism or interruptions in trade. The proportion of plant species that were adventive was 4% in the earliest period, then rose to 13% with improved transportation.  

A large proportion of the native species offered in the late 18th and early 19th centuries were grown for export to Europe (think John Bartram).

Rhodendron maximum; sent to Europe by John Bartram (& invasive in Great Britain and Ireland!); photo from Pl@ntNet.identify

Throughout the 19th and 20th centuries, plants from new regions of the world with unique genetic lineages became increasingly available. Until the mid-19th Century, most non-native plants came from Europe and Eurasia. Beginning in 1850, plants native to temperate Asia composed an increasing percentage of non-native nursery flora. In the period 1900 – 1924, 19% of the ornamental nursery flora originated from temperate Asia. By the next period, 1925 – 1946, this percentage rose to 20.8%.At the same time, North American species (including some from Mexico, Canada, or Alaska) composed 21.9% of the nursery flora. (see graph).

% of species from various origins; North America – medium blue; temperate Asia – dark pink; Europe – tan; Eurasia – fuscia; Southern America – blue-green; Africa – yellow-green; Americas – olive

Plants from East Asia were particularly desirable for both biological and social reasons. First, because of climatic similarities between the two regions, East Asian plants thrived in the eastern United States. Second, popular ornamental genera had higher species richness in East Asia. Important social or cultural factors were a growing fascination with Japanese and Chinese-style gardens: forced “opening” of access to those countries in the 1840s and 1850s; and plant collecting expeditions sponsored by British and American institutions and private collectors. In 1898, the US Department of Agriculture established the Section of Seed & Plant Introduction; its purpose was to collect and cultivate economically useful non-native plants from throughout South America and Asia.

As I noted above, diversity of species in nursery offerings reached a peak in the first years of the 20th Century, concurrent with the first wave of US-sponsored plant collections; indeed, 70 species that were first listed after 1911 in their dataset were introduced by the USDA introduction program between 1912 -1942.

Commodore Perry in Japan; Library of Congress

Counter-pressures and Counter-measures

There were counter pressures during this period that – as mentioned above—probably contributed to a decline in plant introductions in later years. In the 1890s, several US states began requiring inspection of imported plant materials (spurred by plant disease outbreaks caused by spread of San Jose scale from California).

Congress adopted the Plant Quarantine Act in 1912; USDA implemented it through stringent regulations issued in 1919 (Quarantine-37). I have already noted interruption of trade associated with WWI and WWII. Kinlock et al. don’t mention the Great Depression that intervened, but I think it played a role, too. On the other hand, Q-37 was relaxed to target particular species or regions based on pest risk analysis. The article says the relaxation began in the 1930s, but I believe it actually was during the 1970s; see Liebhold et al. 2012. I have blogged several times about how well the current regulations – including the “NAPPRA” program – prevent introductions of invasive plants or damaging plant pests. To review these blogs, scroll down below the archives to the “categories” section and click on “plants as pest vectors”.

dogwood anthracnose; photo by Robert Anderson, USFS; via Bugwood

SOURCES

Adams, D.W. 2004. Restoring American Gardens: An encyclopedia of heirloom ornamental plants. Timber Press

Kinlock, N.L., D.W. Adams, M. van Kleunen. 2022. An ecological and evolutionary perspective of the historical US nursery flora. Plants People Planet. 2022;1–14. wileyonlinelibrary.com/journal/ppp3

Liebhold, A.M., E.G. Brockerhoff, L.J. Garrett, J.L. Parke, and K.O. Britton. 2012. Live Plant Imports: the Major Pathway for Forest Insect and Pathogen Invasions of the US. www.frontiersinecology.org

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

www.fadingforests.org

Plant Invasions’ Impacts on Wildlife

spotted knapweed (Centaurea maculosa); photo by Alan Vernon via Wikipedia

Litt and Pearson (full citation at the end of the blog) are trying to improve scientists’ ability to predict the impact of biological invasions. Their goal is to predict which organisms will be winners, which losers, in the face of anthropogenic ecosystem change.

They focus on exotic plant invasions, because they are ubiquitous. Furthermore, plant invasions affect ecosystems by reassembling the plant community in ways that affect the niches used by native animals and hence the animals’ success under the new conditions. After determining the differences between the traits exhibited by invasive plants vs. the native plants they are displacing, scientists can then identify which native animals are most likely to be affected, as well as how and why they might respond to exotic plant invasion. [Note that Doug Tallamy is looking at similar issues.]

Litt and Pearson have developed a framework to assess how plants’ traits might affect associated wildlife. Applying the framework requires certain baseline information about the ecosystem in question.

This knowledge is applied in stepwise fashion:

1) Identify the fauna of interest and their linkage to the native plant community. This association might be food or habitat values such as shelter. Then the researcher determines the relevant plant traits of importance to that animal and approximates the strength of the animal’s dependence on these traits. Note that the focus is on plant traits relevant to the animal users, rather than specific plant species.

2) Determine overall importance of the plant traits for the area under study by (a) averaging dependence of a representative subsample of individuals to obtain a community-level value for each plant species or functional group and (b) quantifying the relative abundance of the plant functional group in the community (e.g., cover or biomass).

3) Plot the way the animal species’ abundance changes with resource abundance.

4) Understand how the invasive plants will alter the distributions of the native plants’ traits and potentially introducing novel traits that might alter the faunal community.

Litt and Pearson reviewed earlier studies to test how well this framework explained the responses of three groups of fauna to plant invasions in different ecosystems.

searching for spotted knapweed; photo by Oregon Department of Agriculture

Spiders in invaded grasslands

Intermountain grasslands of western Montana are heavily invaded; non-native plants already comprise 25–60% of average total plant cover.

One group of native spiders construct their irregular webs entirely within a single plant. A second group – orb weavers – suspend their larger webs from multiple plants. The former depend on the architectural complexity of individual plants; they can build larger webs in plant species possessing greater branching and/or longer branches of the flowering stalks. Orb spiders depend more on the complexity of the overall plant community.

Plant architecture is closely tied to the plant’s functional groups, that is, whether they are grasses or forbs.

These grasslands are generally dominated by perennial grasses. The irregular-web spiders can use grasses, but strongly favor forbs, particularly those with the most complex flowering structures. Orb weavers are generalists, incorporating multiple plant species; but they also tend to favor forbs, presumably because they are more robust.

Invasive plants in the Western Montana grasslands are of two types: an annual grass, cheatgrass (Bromus tectorum), and numerous perennial and annual forbs. Cheatgrass largely replaces the dominant native grasses with a similar architecture – although cheat is shorter. The exotic forbs, which can collectively invade at levels comparable to cheatgrass, tend to be taller and more complex structurally than the native forbs. Thus, invasion by exotic forbs strongly shifts the community-level distribution of the key trait toward greater structural complexity by replacing the dominant, but structurally simplistic, native grasses, and the more diminutive native forbs. These changes increased the abundance of both spider groups, but especially the specialist irregular web weavers. They find the new conditions meet their needs. Both spider groups appeared to expand their realized niches in response to invasion, i.e., they are able to use a broader range of plant architectures than was available in the native system.

Chaetodipus sp. photo by J.N. Stuart

Rodents in semi-desert grasslands invaded by Lehmann lovegrass

In the semi-desert grasslands of the American southwest, native grasses and forbs provide food and habitat for a variety of rodents. This vegetation influences which species of rodents are present in two ways: the size of the plants’ seeds and the density of vegetative cover. Litt and Steidl examined both. They divided the rodents into separate guilds based on diet and preferred vegetative cover. The two sets of guilds did not overlap for all species.

In southern Arizona, the native plant community is dominated by several grass species and herbaceous forbs; most species produce relatively large seeds. Vegetative cover is generally low, but varies in a patchy fashion. The rodent communities in uninvaded native grasslands are dominated by seed-eaters that prefer sparse cover.

Invasion of these grasslands by Lehmann lovegrass (Eragrostis lehmanniana) results in increased vegetative cover but the grass produces very small seeds that probably provide little to no food for  rodents. Another result is a decrease in overall abundance of arthropods. The new conditions favor different rodent species from those most common in uninvaded habitat.

Two more specialized seed-eating rodent species, which seek both lower cover and larger seeds, decreased in abundance. A rodent species which favors lower vegetative cover and feeds on larger invertebrates also declined. In contrast, abundance increased for two other rodent species that prefer more dense cover and are more opportunistic in their feeding. One species surprised the scientists: Dipodomys merriami increased in abundance, despite the fact that this species favors more open environments. Perhaps other functional traits or biotic interactions are important to this species? There was no apparent change in abundance for three other species, suggesting either a lack of statistical power (2 were less abundant) or that these rodents were able to persist through a balance of positive and negative changes in food and habitat characteristics.

Lucy’s warbler [nest in saguaro, not cottonwood); photo by Dominic Sherony

Warblers in Riparian Habitats in the Southwest

Riparian habitats in the same desert region have been aggressively invaded by the exotic shrub saltcedar (Tamarix spp.). Litt and Pearson consider the findings of Mahoney et al. of this invasion’s impact on two ecologically similar warbler species. One, the yellow warbler (Setophaga petechia), is very widely distributed across North America; it is considered a generalist. The other, Lucy’s warbler (Oreothlypis luciae), is endemic to a small region of the southwest United States and northern Mexico.

The two species have similar feeding behaviors but differ in their nesting requirements. The yellow warbler constructs open cup nests in the branches of shrubs and trees. Lucy’s warbler nests in cavities in larger trees excavated by others. Hence, these species were expected to respond similarly to changes in food resources and foraging habitat, but differ in their responses to changes in nesting substrate.

Native vegetation in the region consists primarily of willows and cottonwoods in the riparian corridors, with oak and mesquite woodlands in the adjacent uplands. Saltcedar invasion rapidly displaces the willows; it takes much longer to displace cottonwoods since are large and long-lived. Upland vegetation is uninvaded and unaffected. While saltcedar is structurally similar to native willows, its leaf architecture allows more light to penetrate in saltcedar stands. This can exacerbate heat stress on nestlings in these hot, arid environments, as well as expose the nestlings to nest predation. These effects are exacerbated by the presence of a biocontrol leaf beetle (Diorhabda spp.), which cause widespread defoliation of saltcedar during nesting season. Meantime, the cavity nests used by Lucy’s warbler are barely affected.

The study by Mahoney et al. showed that in low-invasion riparian sites, the two warblers occur at comparable abundances. When saltcedar invasion replaces willows, yellow warblers decline by ~50% while there is no apparent change in abundance of Lucy’s warblers.

Litt and Pearson point out that their framework is based on two key assumptions that establish the context for its efficacy.

The first is that bottom-up forces fuel ecological processes. Plants are key to making the sun’s energy available to consumer animals and – thence to predators. Consumers’ and predators’ top-down effects are secondary. The authors’ framework thus provides better predictions of community outcomes when systems are predominantly structured by bottom-up forces. As top-down forces increase or when invasive plants differentially affect multiple dimensions of the consumer niche space, it will be more challenging to track and predict outcomes, as our rodent example demonstrates.

The second assumption is that exotic plant invasions will most strongly influence bottom-up processes. Invasive plants displace native plants and their plant traits, thus directly affecting consumers by altering the quality and quantity of food and habitat resources. However, plant community changes caused by plant invasions can also affect predators directly and indirectly via several interactions. These changes in predators’ abundance and/or their per capita effects on prey might create feedbacks that can complicate interpreting and predicting invasion outcomes.

Litt and Pearson concluded that their approach is promising but has inherent limitations linked to the dynamic nature of ecological systems.

[Ecologists continue to evaluate the impacts of saltcedar eradication efforts on another bird species, the federally endangered southwestern willow flycatcher (Empidonax extimus trailii). See, for example, Goetz, A., I. Moffit and A.A. Sher. 2022. Recovery of a native tree following removal of an invasive competitor with implications for endangered bird habitat. Biological Invasions Vol. 24, pp. 2769-2793.]

SOURCE

Litt, A.R. and D.E. Pearson. 2022. A functional ecology framework for understanding and predicting animal responses to plant invasion. Biol Invasions   https://doi.org/10.1007/s10530-022-02813-7 

& Supporting Information [warblers in riparian ecosystems invaded by tamarisk]

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

www.fadingforests.org

Invasive Species Costs Point to Inadequate Effort – especially Prevention

EAB-killed ash tree falls before it can be taken down; photo courtesy of former Ann Arbor mayor John Hieftje

Concerned by growing impacts of bioinvasion and inadequate responses by national governments worldwide and by international bodies, a group of experts have attempted to determine how much invasive species are costing. They’ve built the global database – InvaCost. See Daigne et al. 2020 here.

Several studies have been based on these data. In two earlier blogs, I summarized two of these articles, e.g., Cuthbert et al. on bioinvasion costs, generally, and Moodley et al. on invasive species costs in protected areas, specifically. Here, I look at two additional studies. Ahmed et al. focusses on the “worst” 100 invasives affecting conservation — as determined by the International Union of Conservation and Nature (IUCN). The second, by Turbelin et al., examines pathways of introduction. Full citations of all sources appear at the end of this blog.

It is clear from all of these papers that the authors (and I!) are frustrated by the laxity with which virtually all governments respond to bioinvasions. Thus more robust actions are needed. The authors and I also agree that data on economic costs influence political decision-makers more than ecological concerns. However, InvaCost – while the best source in existence — is not yet comprehensive enough to generate the thoroughly-documented economic data about specific aspects of bioinvasion that would be most useful in supporting proposed strategies.

Scientists working with InvaCost recognize that the data are patchy. At the top level, these data demonstrate high losses and management costs imposed by bioinvasion. The global total – including both realized damage and management costs – is estimated at about $1.5 trillion since 1960. In fact, these overall costs are probably substantially underestimates (Cathbert et al.). [For a summary of data gaps, go to the end of the blog.] Furthermore, they recognize that species imposing the highest economic costs might not cause the greatest ecological harm (Moodley et al).

citrus longhorned beetle exit hole in bonsai tree; USDA APHIS photo

Comparing estimated management costs to estimated damage, the authors conclude that countries invest too little in bioinvasion management efforts and — furthermore — that expenditures are squandered on the wrong “end” of bioinvasion – after introduction and even establishment, rather than in preventive efforts or rapid response upon initial detection of an invader. While I think this is true, these findings might be skewed by the fact that fewer than a third of countries reporting invasive species costs included data on specifically preventive actions. Cuthbert et al. notes that failing to try to prevent introductions imposes an avoidable burden on resource management agencies. Ahmed et al. developed a model they hope will overcome the perverse   incentives that lead decision-makers to either do nothing or delay.

  1. Why Decision-Makers Delay

Citing the InvaCost data, the participating experts reiterate the long-standing call for prioritizing investments at the earliest possible invasion stage. Ahmed et al. found that this was the most effective practice even when costs accrue slowly. They ask, then, why decision-makers often delay initiating management. I welcome this attention because we need to find ways to rectify this situation.

They conclude, first, that invasive species threats compete for resources with other threats to agriculture and natural systems. Second, Cuthbert et al. and Ahmed et al. both note that decision-makers find it difficult to justify expenditures before impacts are obvious and/or stakeholders demand action. By that time, of course, management of invasions are extremely difficult and expensive – if possible at all. I appreciate the wording in Ahmed et al.: bioinvasion costs can be deceitfully slow to accrue, so policy makers don’t appreciate the urgency of taking action.

Cuthbert et al. also note that impacts are often imposed on other sectors, or in different regions, than those focused on by the decision-makers. Stakeholders’ perceptions of whether an introduced species is causing a “detrimental” impact also vary. Finally, when efficient proactive management succeeds – prevents any impact – it paradoxically undermines evidence of the value of this action!

Ahmed et al. point out that in many cases, biosecurity measures and other proactive approaches are even more cost effective when several species are managed simultaneously. They cite as examples airport quarantine and interception programs; Check Clean Dry campaigns encouraging boaters to avoid moving mussels and weeds; ballast water treatment systems; and transport legislation e.g., the international standard for wood packaging (ISPM#15) [I have often discussed the weaknesses in ISPM#15 implementation; go to “wood packaging” under “Categories” (below the archive list)].

pallet “graveyard”; photo by Anand Prasad
  • Pathways of Species’ Introduction

Tuberlin et al. focus on pathways of introduction, which they say influence the numbers of invaders, the frequency of their arrival, and the geography of their eventual distribution. This study found sufficient data to analyze arrival pathways of 478 species – just 0.03% of the ~14,000 species in the full database. They found that intentional pathways – especially what they categorized as “Escape” – were responsible for the largest number of invasive species (>40% of total). On the other hand, the two unintentional pathways called “Stowaway” and “Contaminant” introduced the species causing the highest economic costs.

Tuberlin et al. therefore emphasize the importance of managing these unintentional pathways. Also, climate change and emerging shipping technologies will increase potential invaders’ survivability during transit. Management strategies thus must be adapted to countering these additive trends. They suggest specifically:

  • eDNA detection techniques;
  • Stricter enforcement of ISPM#15 and exploring use of recyclable plastic pallets (e.g., IKEA’s OptiLedge); [see my blog re: plastic pallets, here]
  • Application of fouling-resistant paints to ship hulls;
  • Prompt adoption of international agreements addressing pathways (they cite the Ballast Water Management Treaty as entered into force only in 2017 — 13 years after adoption);
  • Ensuring ‘pest free status’ (per ISPM#10) before allowing export of goods—especially goods in the “Agriculture”, “Horticulture”, and “Ornamental” trades; and
  • Increasing training of interception staff at ports.

What InvaCost Data say re: Taxa of greatest concern to me

Two-thirds of reported expenditures are spent on terrestrial species (Cuthbert et al.). Insects as a Class constitute the highest number of species introduced as ‘Contaminants’ (n = 74) and ‘Stowaways’ (n = 43). They also impose the highest costs among species using these pathways. Forest insects and pathogens account for less than 1% of the records in the InvaCost database, but constitute 25% of total annual costs ($43.4 billion) (Williams et al., in prep.). Indeed, one of 10 species for which reported spending on post-invasion management is highest is the infamous Asian longhorned beetle (Tuberlin et al.)

ALB pupa in wood packaging; Pennsylvania Dept. of Natural Resources via Bugwood

Mammals and plants are often introduced deliberately – either as intentional releases or as escapes. Plant invasions are reported as numerous but impose lower costs.

Tuberlin et al. state that intentional releases and escapes should in theory be more straightforward to monitor and control, so less costly. They propose two theories: 1) Eradication campaigns are more likely to succeed for plants introduced for cultivation and subsequently escaped, than for plants introduced through unintentional pathways in semi-natural environments. 2) Species introduced unintentionally may be able to spread undetected for longer; they expect that better measures already exist to control invasions by deliberate introductions. I question both. Their theories ignore that constituencies probably like the introduced plants … and the near absence of attention to the possible need to control their spread. This is odd because elsewhere they recognize conflicts over whether to control or eradicate “charismatic” species.

Geographies of greatest concern to me

  • North America reported spending 54% of the total expenditure in InvaCost. Oceania spent 30%. The remaining regions each spent less than $5 billion. (Cuthbert et al.)
  • North America funded preventative actions most generously than other regions. Cuthbert suggests this was because David Pimentel published an early estimate of invasive species costs. I doubt it. The Lacey Act was adopted in 1905. USDA APHIS was formed in 1972 – based on predecessor agencies — because officials recognized the damage by non-native pests to agriculture. APHIS began addressing natural area pests with discovery of the Asian longhorned beetle in 1996. Of course, most of APHIS’ budget is still allocated to agricultural pests. I conclude that North America’s lead in this area has not resulted in adequate prevention programs.
Oregon ash swamp before attack by EAB (photo by Wyatt Williams, Oregon Dept. of Forestry)

Equity Issues

Tuberlin et al and Moodley et al. address equity issues of who causes introductions vs. who is impacted. This is long overdue.

  • More than 80% of bioinvasion management costs in protected areas fell on governmental services and/or official organizations (e.g. conservation agencies, forest services, or associations). With the partial exception of the agricultural sector, the economic sectors that contribute the most to movement of invasive species are spared from carrying the resulting costs (Moodley et al.)
  • A lack of willingness to invest might represent a moral problem when the invader’s impacts are incurred by regions, sectors, or generations other than those that on whom management action falls (Ahmed et al.)
  • People are perhaps more inclined to spend money to mitigate impacts that cause economic losses than those that damage ecosystems (Tuberlin et al.)

Data deficiencies

  • Only 41% of countries (83 out of 204) reported management costs; of those, only 24 reported costs specifically associated with pre-invasion (prevention) efforts (Cuthbert et al.).
  • Reliable economic cost estimates were available for only 60% of the “worst” invasive species (Cuthbert et al.)
  • Only 55 out of 266,561 protected areas reported losses or management costs (Moodley et al.).
  • Information on pathways of introduction was available for only three species out of 10,000 (Turbelin et al).
  • Taxonomic and geographic biases in reporting skew examples and possibly conclusions (Cuthbert et al.).

SOURCES

Ahmed, D.A., E.J. Hudgins, R.N. Cuthbert, .M. Kourantidou, C. Diagne, P.J. Haubrock, B. Leung, C. Liu, B. Leroy, S. Petrovskii, A. Beidas, F. Courchamp. 2022. Managing biological invasions: the cost of inaction. Biol Invasions (2022) 24:1927–1946 https://doi.org/10.1007/s10530-022-02755-0

Cuthbert, R.N., C. Diagne, E.J. Hudgins, A. Turbelin, D.A. Ahmed, C. Albert, T.W. Bodey, E. Briski, F. Essl, P. J. Haubrock, R.E. Gozlan, N. Kirichenko, M. Kourantidou, A.M. Kramer, F. Courchamp. 2022. Bioinvasion costs reveal insufficient proactive management worldwide. Science of The Total Environment Volume 819, 1 May 2022, 153404

Moodley, D., E. Angulo, R.N. Cuthbert, B. Leung, A. Turbelin, A. Novoa, M. Kourantidou, G. Heringer, P.J. Haubrock, D. Renault, M. Robuchon, J. Fantle-Lepczyk, F. Courchamp, C. Diagne. 2022. Surprisingly high economic costs of bioinvasions in protected areas. Biol Invasions. https://doi.org/10.1007/s10530-022-02732-7

Turbelin, A.J., C. Diagne, E.J. Hudgins, D. Moodley, M. Kourantidou, A. Novoa, P.J. Haubrock, C. Bernery, R.E. Gozlan, R.A. Francis, F. Courchamp. 2022. Introduction pathways of economically costly invasive alien species. Biol Invasions (2022) 24:2061–2079 https://doi.org/10.1007/s10530-022-02796-5

Williams, G.M., M.D. Ginzel, Z. Ma, D.C. Adams, F.T. Campbell, G.M. Lovett, M. Belén Pildain, K.F. Raffa, K.J.K. Gandhi, A. Santini, R.A. Sniezko, M.J. Wingfield, and P. Bonello 2022. The Global Forest Health Crisis: A Public Good Social Dilemma in Need of International Collective Action. Submitted

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

www.fadingforests.org

Search for Asian giant hornet

Asian giant hornet (Vespa mandarinia); photo by University of Florida Dept. of Entomology

Washington State’s “Giant Hornet – Hornet Herald” for June asks people to help with detecting this pest by monitoring paper wasp nests (hornets attack them). Hornet visits last 5 – 10 minutes while the hornet removes paper wasp larvae.  How to help:

  • Locate paper wasp nests that you have access to and can monitor through October. Then log the nest locations using the form here
  • Visit the nests each week, observe them, and then log your nest activity on a different form – here. Please monitor the nests for at least 5 minutes during the day once per week, but you can check the nests for as long and as often as you would like.

If you would like guidance on how to become a citizen-science monitor or trapper of Asian giant hornets – or presumably other bioinvaders – go here

Meanwhile, Washington State Department of Agriculture entomologists are in South Korea testing several hornet attractants and studying hornet foraging behavior. The goal  is to improve Washington’s trapping and tracking techniques.

Of course, 2022 is only half over, but so far neither Washington nor British Columbia has confirmed any detections.

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

www.fadingforests.org