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Save A Place at the Table: Is There a Place for Non-Natives in Ecological Restoration?

By Andrea Gregor

Non-native species are known to be a strong driver of native species decline and habitat degradation (D’Antonio & Meyerson, 2002). All invasive species are non-native, yet not all non-native species are invasive (Clewell & Aronson, 2013). Non-native species have the ability to infect natives with disease, outcompete them and alter ecosystem functions (D’Antonio & Meyerson, 2002). Non-native species, however, have also been shown to enhance the process of ecological restoration by acting as an alternate food source, or increasing nutrients in soil and becoming an important part of ecosystems. I focus on non-native plant species, however non-native animal species follow similar trends, and equal research should be performed. If ecological restoration is the restoring of native landscapes, then is there any room for non-native species to be part of this process? How can we assure that non-natives used in ecological restoration will not become invasive?

The Potential Impact of Non-Native Species

An area of land that is to go through ecological restoration often has had disturbance caused by human action or environmental events (Keenleyside, Dudley, Cairns, Hall, & Stolton, 2012). According to Vilà and Weiner (2004), disturbance increases the chance of invasion by non-native species as species that have the potential to become invasive tend to be good colonisers after disturbances. As well as this, many create seed banks which allow them to endure for a long period, and make eradication difficult. With an increased risk of invasion, we get an increased risk of interspecific competition between native and non-native species. Studies conducted are unable to say that all non-native species always outcompete natives; however there is still a strong competitive effect on native species, which can cause a decline in the population of native species (Vilà & Weiner, 2004). Species compete for light, nutrients and space (Wilson & Tilman, 1993). Therefore including certain species of non-natives in restoration runs the risk that the introduction, be it accidental or not, could be detrimental to the persistence of that ecosystem through potential outcompeting and overcrowding.

If inadequate research is done, non-native species have the potential to become invasive in certain environments. Invasive species have been recognised as the second largest threat to global biodiversity after habitat fragmentation (Allendorf & Lundquist, 2003). Throughout the world, invasive species cost governments billions of dollars. Management of plants and animals listed under the Endangered Species Act cost $32-$42 million annually, in which 90% of those funds are allocated to mitigate the effects of invasive species (Wilcove & Chen 1998; D’Antonio & Meyerson 2002). In New Zealand invasive species cost $840 million each year to control, and produce a $1 billion loss in productivity (Giera & Bell, 2009). With such a large economic impact that invasive species have on New Zealand and the world, should we risk using species that have the potential to be invasive in ecological restoration?

Are All Non-Natives That Evil?

hummingbird-bergez-02

Figure 1: American hummingbird feeding on honeysuckle. John Bergez 2012

With adequate research, there is room to include non-natives in ecological restoration. Some non-native species, particularly plant species have been shown to increase the population of native species. Non-native species being used as an alternate food source for native species can lead to an increase in native population numbers due to the increased resources. For example, in the US, introduced honeysuckles are improving native bird populations (Figure 1). It is also found that seed dispersal of native plants is the highest where non-native honeysuckles are the most abundant due to dispersal by the now more populated native birds (Davis, et al., 2011). This positive effect of a non-native plant has enhanced the population of native bird species, as well as other native plants. Subsequently, removing non-native species can have negative effects to an ecosystem removal of these pine plantations will demolish the favourable , and successful eradication so far has been limited to small islands (Zavaleta, Hobbs, & Mooney, 2001). With declining native habitat, half of New Zealand’s threatened indigenous plants are found in historically rare ecosystems with localised distributions (Pawson, Ecroyd, Seaton, Shaw, & Brockerhoff, 2010). Encouraging natives to use non-native habitats as substitutes could help the continuation of species. The New Zealand large bird orchid (Chiloglottis valida) has been found within non-native Pinus nigra plantations. The microclimate under these pine trees which allow orchids to survive outside of their original habitat. Because of this reason, a small orchid reserve in this plantation has been created, while the rest of the plantation has been logged (Pawson, Ecroyd, Seaton, Shaw, & Brockerhoff, 2010). This example is one of many which show the necessity of keeping specific non-native species in order to retain native species.

Implications on Soil Nutrients

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Figure 2: The soil alternating Falcataria moluccana. HELCO 2014 

Exotic plants can alter ecosystem processes through differences in nutrient cycles. They can cause an increase, or decrease to the soil nutrients created by native species (Ehrenfeld, 2003). Negative effects of the changes in the soil microbial community can lead to an increase in the invasiveness of an ecosystem from other species (Green, O’Dowd, Abbot, Jeffery, Retallick, & Mac Nally, 2011). This has the potential to create an invasion meltdown which would undo all restoration efforts thus far and render a project useless. Changes in soil structure, such as an increase in nitrogen can encourage growth from other invasive species which outcompete natives, and shroud out the light with denser canopies, reducing the growth of native species (Allison, Nielsen, & Hughes, 2006). In Hawai’i, nitrogen fixing invasive tree Falcataria moluccana (figure 2) alters the soil structure which limits the growth of native species. Along with this, F. moluccana facilitates the invasion of another non-native species, Psidium cattleianum (also known as the strawberry guava) which outcompete natives for resources (Allison, Nielsen & Hughes, 2006). If a species such as this was used in ecological restoration without research, it has the potential to become invasive and harm the ecosystem, rather than benefit it. Changes in soil structure can also have positive effects; non-native species are able to be used positively as substitutes for slower growing natives when restoring areas with poor productivity soils which have had disturbances such as overgrazing or mining (Wong, 2003). Fast growing nitrogen fixing trees from Asia were found to grow well in degraded pastures in Puerto Rico and accelerated regeneration of native forests (D’Antonio & Meyerson, 2002). Without exotic species in circumstances like this, ecological restoration would not be able to get under way, especially if natives  we are wanting to maintain struggle to establish in degraded soils. Non-natives in this case are essential to effective recovery of native sites. The differences in the nutrient cycles of non-native species and native species we want to restore will determine the impact non-natives will have on the soil composition and therefore the native species. These impacts can differ from site to site and can cause ecological restoration to fail or succeed. Research is our greatest tool to ensure we only use species that succeed, and remove species that will cause our project to fail.

 

Novel Ecosystems

With continued human movement, non-natives are becoming more and more abundant, creating many novel ecosystems (Marris, 2011). In ecological restoration, you must pick your battles; it is not possible to remove all introduced species. If a non-native has no potential of becoming invasive, and is doing no harm to an ecosystem, then leaving that species and focusing money on other areas would seem like the way forward. In some cases, I feel we should learn to embrace novel ecosystems, especially in circumstances where we are unable to return ecosystems back to their original state. Non-natives may have changed the habitat of an area to make it unsuitable for future natives, whether the non-native is present or not through changes in soil or species composition, and abundance (Norton, 2009) . According to Norton (2009), it has passed the biotic threshold, and there is no way to return the ecosystem back to its original state. If this is to have happened, and a non-native has taken over the niche of a native without affecting other species, it may be in our best interests just to embrace the change leave it there as part of a functioning ecosystem.

Conclusion

Non-native species have a place in ecological restoration, however we must be wary of which species we choose to include in these projects. The fact that invasive non-native species are one of the largest threats to ecological restoration means that using non-natives in these practises can lead us to walk on a fine edge between enhancing native species, and causing an invasive meltdown. Introducing non-native species seems like we are encouraging the opposite of what we are trying to achieve, however it has been shown in many cases to work. We have little room for error, therefore we must use short lived, well researched species and we must monitor them closely to ensure ecological restoration is achieved successfully. We must also acknowledge that there is always a chance of failure; a species may interact with its surrounding different than we had planned. This is a risk that is shared in all conservation and restoration projects which can be minimised, but never removed. In this essay.  We must also look at the possibility of leaving non-natives that have been determined low risk to ecosystems; we can never restore every area back to its original state, but if we pick our fights correctly, we are able to nurse many native species back with the help of non-natives.

 

Bibliography

Allendorf, F. W., & Lundquist, L. L. (2003). Introduction: Population Biology, Evolution, and Control of Invasive Species. Conservation Biology Vol. 17 (1), 24-30.

Allison, S., Nielsen, C., & Hughes, R. (2006). Elevated enzyme activities in soils under the invasive nitrogen-fixing tree Falcataria moluccana. Soil Biology and Biochemistry Vol. 38(7), 1537-1544.

Clewell, A. F., & Aronson, J. (2013). Ecological Restoration – Principles, Values & Structure of an Emerging Profession (2nd ed.). Washington, D.C: Island Press.

D’Antonio, C., & Meyerson, L. (2002). Exotic Plant Species as Problems and Solutions in Ecological Restoration: A synthesis. Restoration Ecology Vol 10 (4), 703-713.

Davis, M. A., Chew, M. K., Hobbs, R. J., Lugo, A. E., Ewel, J. J., Vermeij, G. J., et al. (2011). Don’t judge species on their origins. Nature Vol 474, 153-154.

Ehrenfeld, J. G. (2003). Effects of Exotic Plant Invasions on Soil Nutrient Cycling Processes. Ecosystems Vol. 6 (6), 503-523.

Forbes, A. S., Norton, D. A., & Carswell, F. E. (2015). Underplanting degraded exotic Pinus with indigenous conifers assists forest restoration. Ecological Management and Restoration Vol 16(1), 41-49.

Giera, N., & Bell, B. (2009). Economic Costs of Pests to New Zealand. Wellington: Crown Copyright- Ministry of Agriculture and Forestry.

Green, P. T., O’Dowd, D. J., Abbot, K. L., Jeffery, M., Retallick, K., & Mac Nally, R. (2011). Invasional meltdown: Invader–invader mutualism facilitatesa secondary invasion. Ecology Vol 92(9), 1758-1768.

Keenleyside, K., Dudley, N., Cairns, S., Hall, C., & Stolton, S. (2012). Ecological Restoration for Protected Areas-Principles, Guidelines and Best Practices. Gland: International Union for Conservation of Nature and Natural Resources.

Marris, E. (2011). Rambunctious Garden. New York: Bloomsbury.

Norton, D. A. (2009). Species Invasions and the Limits to Restoration: Learning from the New Zealand Experience. Science Vol 325 (5940), 569-571.

Pawson, S. M., Ecroyd, C. E., Seaton, R., Shaw, W. B., & Brockerhoff, E. G. (2010). New Zealand’s exotic plantation forests as habitats for threatened indigenous species. New Zealand Journal of Ecology Vol 34 (3), 342-355.

Schlaepfer, M. A., Sax, F. D., & Olden, J. D. (2011). The Potential Conservation Value of Non-Native Species. Conservation Biology Vol. 25 (3), 428-437.

Vilà, M., & Weiner, J. (2004). Are invasive plant species better competitors than native plant species? – evidence from pair-wise experiments. OIKOS Vol 105(2), 229-238.

Wilson, S. D., & Tilman, D. (1993). Plant Competition and Resource Availability in Response to Disturbance and Fertilization. Ecology Vol 74(2), 599-611.

Wong, M. H. (2003). Ecological Restoration of mine degraded soils, with emphasis on metal contaminated soils. Chemosphere Vol 50, 775-780.

Zavaleta, E. S., Hobbs, R. J., & Mooney, H. A. (2001). Viewing invasive species removal in a whole-ecosystem context. TRENDS in Ecology & Evolution Vol 16 (8), 454-459.

 

 


A New Way to Combat Introduced Diseases: The Use of Synthetic Biology in Conservation – By Feyza Zaman

The current ecological atmosphere is arguably one of instability, over-exploitation, and a rapid and poorly understood change towards global homogeneity. The numerous anthropogenic pressures placed on the natural environment, potentially beginning with early humans millennia ago, have been ever increasing in volume and effect, with extinction rates now thought to be approximately 100-1000 times greater than what is considered normal background levels (Smith et al. 2009). These ongoing impacts are now thought to be leading towards a sixth mass extinction of biota on earth, comparable in magnitude to the Cretaceous-Palaeogene extinction which wiped out the dinosaurs (Raup and Sepkoski 1982; Barnosky et al. 2011). Currently, the negative impacts which human activities have upon biodiversity can be broadly placed into four categories; habitat destruction, over-exploitation, the introduction of non-native species, and the subsequent spread of diseases caused by pathogens carried by these species (Wilcove et al. 1998; Smith et al. 2009). Although the first three of these impacts have been found by many studies to have the greatest influence upon the abundance and diversity of native communities (e.g. Venter et al. 2006), it is also acknowledged that the role of introduced diseases is not yet fully understood. However, it is now coming to light that such diseases may pose a much greater risk than previously thought, and will require the development of new methods of treatment and conservation to deal with these threats (Smith et al. 2006; Smith et al. 2008; Fisher et al. 2012).

There has been an apparent increase in the emergence of infectious diseases in wildlife in recent times, and there is substantial evidence that these outbreaks have been greatly affecting local species populations by causing severe and potentially permanent declines in abundance (Harvell et al. 2002; Smith et al. 2006; Smith et al. 2008; Fisher et al. 2012). Factors thought to be contributing to this increase include changes in environmental dynamics, such as increased susceptibility from stress-causing temperature fluctuations; ecological influences, such as reduced genetic diversity from other factors like habitat loss; and socio-economic factors, including globilisation and increasing mobility resulting in more introduced species (Smith et al. 2006; Jones et al. 2008; Smith et al. 2012). There are many pathogens which are known to result in widespread reductions in wildlife diversity, however much attention has recently been placed on several fungal agents of disease which are known to have resulted in the ecological collapse of formerly highly abundant species, and a prominent example of this is blight in the American chestnut (Fisher et al. 2012; Powell 2014). In this review, we will be focusing on the chestnut blight as a case study, outlining the potential processes necessary for advising further action in other disease epidemics which have appeared worldwide.

With these relatively new challenges to biodiversity conservation have come numerous approaches attempting to resolve them, ranging from traditional methods to the application of technologies novel to conservation biology. One such novel partnership of disciplines is that of synthetic biology and conservation biology. Synthetic biology involves the manipulation, design, and construction of chemically synthesised DNA or other biological constructs with the aim to meet “human needs by the creation of organisms with novel or enhanced characteristics” (Presidential Commission for the Study of Bioethical Issues 2010; Redford et al. 2013). In the case of disease control, synthetic biology may allow us to more intimately understand the causal agents, potentially leading to methods of reducing their pathogenicity, or to introduce protective alleles into susceptible species. So, whilst traditional methods are still required to support and supplement conservation, they have clearly not achieved their goal of halting biodiversity loss (Butchart et al. 2010). Thus new technologies such as transgenesis, synthetic biology techniques, and DNA manipulation may provide the next step (Redford et al.2013; Powell 2014). These ideas have only in recent years been brought together with conservation biology, and may have the potential to greatly widen the scope of biodiversity protection.

Natural pre-blight range of the American Chestnut across the eastern USA, exceeding 800,000km2 (adapted from Jacobs 2007).

Natural pre-blight range of the American Chestnut across the eastern USA, exceeding 800,000km2 (adapted from Jacobs 2007).

A prominent example of the application of these novel techniques is in the American chestnut tree (Castanea dentata), whose widespread populations were decimated by an introduced fungal blight (Wheeler and Sederoff 2009; Zhang et al. 2013). Before the early 20th century, a quarter of the hardwood forests of the eastern United States constituted of American chestnut, numbering in the billions (Merkle et al. 2007; Powell 2014). This tree was a keystone species in its ecosystem, contributing immensely to forest structure and complexity, and providing habitat and a stable food source for squirrels, bears, deer, and many other animals (Powell 2014). Deliberate introductions in the early 1900s of the Japanese (Castanea crenata) and Chinese (C. mollissima) chestnut trees brought with them spores of the pathogenic fungus Cryphonectria parasitica, the causal agent of chestnut blight (Barakat et al. 2009; Powell 2014). These trees, having evolved with the fungus, had a natural resistance to it; however the American variant did not, never having been exposed to it. Establishing itself below the bark, the blight spreads filamentous hyphae into the tree and produces oxalic acid, among other toxic substances, which lowers the pH of infected tissue to lethal levels. As the toxins spread, a canker forms around the trunk or limb of the tree, and eventually cuts off extremities from water and nutrients, resulting in death (Anagnostakis 1987). This disease ran rampant through the north-eastern populations of American chestnut, killing more than 3 billion individuals within 50 years (Powell 2014). The American chestnut is now functionally extinct within its natural range, with very few adult trees left in various areas around America, mostly outside this natural distribution. It is important to note that, along with the important ecological roles that this species fulfilled, the American chestnut was also very important economically, providing food, fuel, and fast growing and rot resistant timber (Homles et al. 2009).

Different stages of blight disease in American Chestnut (Taken from Anagnostakis 1982 and Powell 2014).

Different stages of blight disease in American Chestnut (Picture credits: Anagnostakis 1982 and Powell 2014).

The imperative nature of the situation was recognised very early on, and over the decades much research has been directed towards mitigating the effects of the blight. There are currently three strategies underway which appear to be having some measure of success: one utilising ancient horticulture cross- and back-breeding techniques; one targeting the fungus directly through an introduced virus; and one utilising genetic engineering to insert specific strands of DNA into the tree’s genome (Grüenwald 2012; Zhang et al. 2013; Powell 2014). The first method is focussing on hybridising the American chestnut with Asian variants, and further back-breeding with the American species in order to create individuals with blight-resistance, whilst retaining as many of the American species’ traits as possible (Diskin et al. 2006; Hebard 2006; Jacobs 2007). However, this method can be very inexact, and requires many generations to achieve the desired results. Projects using this method have spanned over several decades, and there has been some success with several re-plantings of hybrid trees across America (Diskin et al. 2006; Hebard 2006; Jacobs 2007). However, the genes for resistance from the Asian trees is incompletely dominant, and intensive selection through individual maturation, inoculation with the disease to check for resistance, and subsequent crossing of resistant individuals is required (Powell 2014).

The second method involves targeting the blight fungus directly with hypoviruses (a virus that affects only fungi) from the family Hypoviridae. These hypoviruses have the very specific effect of reduced virulence and sporulation of the blight fungus, but has been found to be almost too efficient, commonly killing the fungus on a single tree before the virus is able to spread to other infected trees (Kazmierczak et al. 1996). Although effective if an infection is caught early, this method requires inoculation of individual trees and can be labour and time intensive.

Using modern technology, the third method is attempting to confer disease resistance directly and in a much more precise manner, targeting specific symptoms of the fungus on the trees. Using the natural DNA transmission capabilities of the Agrobacterium tumefaciens bacteria, scientists at the State University of New York (SUNY) were able to transfer genes from a species of wheat resistant to an oxalic acid producing fungus similar to the chestnut blight, Sclerotinia sclerotorium (Welch et al. 2007; Zhang et al. 2013; Powell 2014). The gene codes for the enzyme oxalate oxidase – which chemically breaks down oxalic acid – appears to be working exceptionally well in the transgenic American chestnut trees. Several generations of the modified trees have been created, with varying levels of oxalate oxidase production within their cells, and several strains appear to even be more resistant to the blight than the Asian species (Zhang et al. 2013; Powell 2014). The team at SUNY has stated that as their research continues they intend to introduce further lines of defence against the fungus to safeguard against potential future adaptation by the pathogen, and include enzymes from various other plants to confer less specific anti-fungal properties (Zhang et al. 2013; Powell 2014).

The American Chestnut's distinctive features, from a mature tree. Picture credit: American Chestnut Foundation

The American Chestnut’s distinctive features, from a mature tree. Picture credit: American Chestnut Foundation

The next hurdle to such technological advancement in conservation biology is not scientific, but social and political acceptance. Although genetically modified crops and products have mostly become commonplace amongst our modern societies – although not completely, with notable anti-genetically modified organism movements – the prospect of genetically altering wildlife with the intention of releasing them may require further consideration (Dana et al. 2012; Redford et al. 2013;Powell 2014). Having achieved their initial goals of creating a disease resistant species, acquiring official permission from the Food and Drug Administration and US Department of Agriculture is seen as the next major obstacle by the team at SUNY. With the American chestnut tree having both an important role in ecosystem processes, and high economic value, it is believed that this approval will be achieved within the next 5 years (Powell 2014).

If such a precedent is set, it may pave the way for much wider applications of this technology in the service of conservation biology. There are similar situations to the American chestnut where entire species or even higher taxonomic orders have been subjected to enormous reductions in abundance due to introduced fungal diseases, including the global epidemic of Chytrid disease in amphibians, and the more recent outbreaks of white-nose syndrome among American bats (Weldon et al. 2004; Frick et al. 2010). Both of these diseases have reached pandemic proportions, and tireless research still has yet to provide a solution. Although fundamentally different from the concepts discussed above, which primarily apply to plant ecology, the overall conclusions can mostly be applied to these amphibian and mammalian taxa. The causative agents of the fungal diseases, as well as their symptoms, have been identified – however effective methods of combatting these remain elusive. When facing global dissemination, such as the Chytrid disease, or up to 95% mortality rates as in the White-nose syndrome, urgent action needs to be taken in order to manage and minimise biodiversity loss (Weldon et al. 2004; Frick et al. 2010; Woodhams et al. 2012). If genes of resistance to such diseases can be identified in alternate organisms, and further research is undertaken in biosynthetic and transgenic techniques, it is just a matter of time before the processes detailed in this discussion become applicable to a wider range of pathogens for conservation related purposes.

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