By Jeff Balland
The recent concept of novel ecosystems has aroused many debates. Novel ecosystems can be defined as new systems where new species combinations and functions that have never interacted historically, occur irreversibly and sustainably (Morse et al. 2014), due to anthropogenic activities, species introduction and climate change (Hobbs et al. 2006; Hobbs et al. 2009). The stage between an ecosystem and a novel ecosystem is called “hybrid ecosystem”, and can be defined by a changing system where a return to previous conditions is still possible before it reaches a tipping point (see Hobbs et al. 2013). Almost 12 years after its introduction (see also Chapin & Starfield 2005), two sides are opposed, whether restoration ecologists should integrate the concept of novel ecosystems into practice or not. I attempt to expose and criticize both of them to see what should be retained about this issue.
Embracing the concept
The proponents of this approach argue that it is more relevant to adapt to climate change, and help ecosystems to keep their functions and services when their communities are unbalanced by changing conditions. As most of existing ecosystems are concerned by changes, “novel ecosystems constitute the new normal” (Marris 2010).
As climate change affects species ranges, migrations and invasions (Parmesan 2006) and because non-indigenous species introduction is one of the biggest causes of native communities changes (natives can be excluded by losing competition) (Clavero & Garcia-Berthou 2005), promoting novel ecosystem management is to say tolerating invasive species (Rodriguez 2006). Indeed, invasive species removal has a real cost for governances. For instance, the removal costs to USA more than 22 billion dollars per year for all invasive species (Pimentel et al. 2005). Is Invasive Non-Native Species (INNS) removal compulsory? Many studies showed that sometimes, removing those species could have unexpected negative impacts on native species and ecosystems so that recovery of native species after their removal is not allowed (see Zavaleta et al. 2001; Ewel & Putz 2004): some INNS have even been described as keystone and engineer species (species playing a crucial role in the ecosystem) (Rodriguez 2006; Sousa & Gutiérrez 2009). For example, an invasive tree in Puerto Rico allows some native plants to settle where there were not able before (Lugo 2004). Considering this, exotic species should not be neglected just because they are non-native (Davis et al. 2011).
By the way, the new concept of assisted migration (translocation of species threatened by climate change into more suitable locations), emerging as a solution to face environmental changes, will permit the creation of novel ecosystems in the areas where species are voluntary introduced (Minteer & Collins, 2010).
Finally, the novel ecosystems approach may allow improving quality of ecosystem services in exploited ecosystems such as plantation forestry or agriculture. In their study, Smaill et al.(2014) showed that the Coast Redwood Sequoia sempervirens matched all the considerations of New-Zealand foresters and could deliver better ecosystem services than the actual most exploited species (Pinus radiata). By the way, the Coastal Redwood is already naturalized in some part of the country (Figure 1).
Yet, many scientists strongly disagree with the novel ecosystems concept. In their critique, Murcia et al.(2014) pointed several oversights of such an approach. First, assuming novel ecosystems are “the new normal” is denying successful stories of restoration and ignoring that many ecosystems are well-preserved. Secondly, it is argued that species responses to climate change are unpredictable on a local or regional scale (the usual restoration scales). Furthermore the thresholds of irreversibility in species combination, namely the tipping points determining whether a hybrid ecosystem may recover to the ancestral one or evolve toward a novel ecosystem, are still difficult if not impossible to identify (Aronson et al. 2014). According to the detractors, such a concept could provide a “license to disturb” for resource exploitation companies, and may reduce the investment in research and restoration projects because they may become unnecessary, as transformation of ecosystems may be accepted. At last, introducing or managing new species combinations, often including INNS, is not worth taking the risk and the precautionary principle should be applied to avoid any aggravation of ecosystems perturbations.
Integration in management
According to Hobbs et al.(2014), novel ecosystem approach in conservation can also be an alternative to classical restoration. In this paper, the authors made a framework on how decisions about ecosystem management should be taken (Figure 2), struggling between different limitations the managers could have in regards of management goals.
However, according to the authors, this framework is theoretic and crucially need further implementation. By the way, decision-making processes may be influenced by the degree of sympathy managers have towards novel ecosystems.
The novel ecosystem concept is a new way of looking at the environment. Integrating it in management practices may allow to use what were threats (for example invasive species) as advantages (ecosystem functioning). It may help to preserve species that are jeopardized by climate change though assisted migration, and ecosystem services of exploited lands may be enhanced by selecting species in regards of their ecological functions. In my opinion, the concept is not ignoring successful stories of restoration, nor it will provide “licence to disturb”, because novel ecosystems are not worth studying to replace conservation but to provide alternative management. However, I agree some new approaches such as assisted migration are uncertain because of unpredictable species responses (to climate change, to new community compositions, etc.). Likewise, the difficulty of identifying the tipping points in hybrid ecosystem is an obstacle to management decisions. But it is definitely worth putting energy in further investigations, because of all the knowledge about ecosystem functioning the discovery of these thresholds would bring. The concept crucially needs implementation even if the principle of precaution regarding the risks should be considered. That is why I strongly believe the concept should be embraced only as an ultimate alternative, when neither sufficient protection (reserves, protection status for species, conservation programs…) nor classical restoration can be done. In that way, the novel ecosystem approach will only provide good overcomes and exciting discoveries.
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Lugo, A.E., 2004. The outcome of alien tree invasions in Puerto Rico. Frontiers in Ecology and the Environment 2, 265–273. doi:10.1890/1540-9295(2004)002[0265:TOOATI]2.0.CO;2
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Truitt, A.M., Granek, E.F., Duveneck, M.J., Goldsmith, K.A., Jordan, M.P., Yazzie, K.C., 2015. What is Novel About Novel Ecosystems: Managing Change in an Ever-Changing World. Environmental Management 55, 1217–1226. doi:10.1007/s00267-015-0465-5
By Olivia Quigan
Restoration is becoming an increasingly useful tool in conservation. We can now bring biodiversity back to an area that has been impaired beyond recognition by human activities, such as logging, damming, or open cast mining. Given that we are restoring more and more ecosystems around the world, does this give us leave to destroy ‘pristine’ habitats in order to exploit them to gain access to resources?
The benefits of destroying habitats in order to access resources are mostly of economic value. When cost-benefit analyses for an open cast mine are done, the only environmental outcomes that are considered are those that can be turned into a monetary value (Abelson, 2015). These are quantified as the physical impacts on the environment and how these impact health and agriculture (Abelson, 2015). This ignores the intrinsic value of unique species, as through losing them we reduce global biodiversity – a value that cannot be measured in currency (Campbell, 2014).
An argument could be made for destruction of habitats with a view to restoration, that given enough planning time, species can be saved before the habitat loss occurs. They could then be returned to the habitat during the restoration process, or found suitable homes elsewhere, that are similar to their current habitat. Translocation can be a valid restoration method, however not without its risks. This method was attempted in 2011. To allow for an open cast mine in New Zealand’s South Island, the unique and endemic Powelliphanta augusta Snails were collected and stored in shipping containers with the ultimate goal of introducing them to nearby forests. This resulted in the deaths of 800 individuals due to a technical failure of the refrigeration unit they were stored in (Vallance, 2011). This attempt was a failure, because even with the remaining snails being translocated, they are not
successfully persisting in their new environment, with death rates at new sites of up to thirty per cent (Morris, 2010). There are many other instances of failed translocations. Analyses of many reptile and amphibian translocations between 1991 and 2006 showed that up to 30 per cent of translocations failed in producing self-sustaining populations (Germano & Bishop, 2008). This rate of failure must force us to come to the conclusion that we do not yet have the knowledge to prevent extinctions in the case of a planned environmental degradation, especially when endemic species are involved, as the risk is often too high to justify needless environmental degradation.
The creation of novel landscapes is an inevitable outcome of the anthropogenic influence on the world. As we remove natural habitats, the areas that replace them won’t be the same; no matter how hard we try to restore them. A study by Lugo, Carlo and Wunderle Jr. (2011) looked at the islands of Puerto Rico and the introduced species there.
The forest cover here dropped from 100 per cent to just six per cent by the 1940s. The restoration of much of the forest has included many introduced species, both plants and animals. The resulting forest was a mixture of both, but the native plant species continue to dominate the forests, with cover of over 80 per cent. Native birds continue to be successful, and forage on both natives and introduced plants. The introduced honeybee appears to have adapted to the phenology of the native plants and is an important pollinator (Lugo, Carlo, & Wunderle Jr., 2012). This indicates that novel habitats created by restoration efforts can be sustained with introduced species, but we must continue to protect the native species to ensure lasting intrinsic value of the ecosystem.
Disturbed habitats are more likely to be susceptible to invading species. These are defined as species which “proliferate and noticeably replace native species,” (Clewell & Aronson, 2013). Invasive species with a more generalist way of life will have an advantage over native species, especially if these species have evolved into a more specialist niche (Clewell & Aronson, 2013). This is even more applicable in island habitats, where animals have evolved with limited predators. Using the land for agriculture or industry changes the scope of the ecosystem, and increases the vulnerability of it to invasions from non-native species (Vitousek, D’Antonio, Loope, Rejmanek, & Westbrooks, 1997). Human modification of environments is a major driver the invasion by non-native species. Logged forests in Thailand that were home to an invasive weed experienced an eight-fold reduction in pollinators visiting native species. The invasive beetle, Coccinella septempunctata, showed higher abundances in agricultural grasslands when compared to non-modified areas (Didham, Tylianakis, Gemmell, Rand, & Ewers, 2007). Due to the precious value of native species, the total destruction of a habitat cannot be justified as this disturbance leads to increased vulnerability to species invasions.
The complete destruction of a habitat will always be detrimental to the species living there. To destroy a habitat for monetary gain is to place a value on the uniqueness of habitats, and deem it less important than the economy. The evidence shows that we are not capable of maintaining the integrity of a habitat if we destroy it completely. Disturbed habitats are more likely to allow invasive species, which decimate native populations. Human attempts at preservation by translocation of species often fail. As we cannot guarantee the safety of our unique species, we cannot justify the destruction of any habitat; regardless of how accomplished we are becoming at restoring them.
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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?
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
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.
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.
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.
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Large-scale rodent control reduces pre- and post-dispersal seed predation of the endangered Hawaiian lobeliad, Cyanea superba subsp. superba (Campanulaceae)- Synopsis by Paul RomanPosted: May 21, 2014
Eradication success down under: Heat treatment of a sunken trawler to kill the invasive seaweed Undaria pinnatifidaPosted: May 18, 2014
The Invasion/Conservation Paradox: What happens when an invasive species is also a threatened species? – Mary F. PaulPosted: May 5, 2014
In conservation, the term non-native tends to evoke a knee-jerk association with the keywords: unwanted, invasion, eradication, pest. Particularly here in New Zealand, the experience with introduced species has been extremely negative, resulting in the loss of many species and the complete disruption of an ecosystem.
The line between introduced species and invasive species has proven difficult to define as (like many terms in ecology) the definition can be ambiguous and open to subjective interpretation (Colautti & MacIsaac, 2004). Some define invaders broadly as a widespread non-indigenous species, whilst others limit the term to species that adversely affect habitats economically and ecologically (Colautti & MacIsaac, 2004; IUCN, 2013).
Invasive species are one of the three main threats to global biodiversity, along with habitat loss and climate change. The introduction of new species to a non-native ecosystem can have devastating flow-on effects throughout the community and can result in both environmental and economic damages.
But what happens when a species is considered invasive in one area of the world but considered threatened in another?
There are several examples of both plant and animal taxa which have been successful invaders of new locations yet are experiencing declines in their native habitat. As climates continue to shift, this may become a more and more frequent occurrence when native areas become less suitable and new climatic envelopes in non-native areas become accessible. Brook trout, Monterey Cypress, and Chinese wattle-necked softshell turtles highlight these paradoxes. All three of these have managed to become established in island settings where introductions can be particularly consequential.These examples also raise questions about ex situ conservation, or practicing conservation through relocation, and weighing the costs of an introduced species and deciding whether or not to preserve them outside their native range.
Brook trout (Salvelinus fontinalis) are a species of char native to eastern North America, inhabiting clear, cool freshwater lakes, rivers, streams, and ponds. While their historic range is limited to the East coast of the United States and Canada, they have been introduced to over 47 countries spanning Europe, South America, Africa and Oceania where they are classified as an invasive and damaging species (ISSG, 2014).
Originally introduced to provide recreational and commercial fishing, not all introductions led to establishment but populations are now present throughout the United States, Europe, and New Zealand (EBTJV, 2013). There are many cases of brook trout out competing native species in introduced regions. Non-native brook trout displace bull trout (Salvelinus confluentus) at high elevations within introduced areas throughout the western United States and Canada (Rieman et al., 2006; Warnock & Rasmussen, 2013).
In their native range extending from southern Canada to South Carolina, habitat fragmentation, invasive species, and climate change are causing declines in brook trout populations. Ironically, non-native fish rank as the largest biological threat to brook trout (EBTJV, 2013). Declines in brook trout in native areas have been observed due to interspecific competition and predation on juveniles by brown trout (Salmo trutta) (Fausch & White, 1981). Increased sedimentation and runoff are likely to be contributing factors to the diminished populations, with higher water temperatures due to industrial runoff and climate change driving the populations to higher elevations. Climate change will continue to restrict the native range of the brook trout, increasing minimum elevation by up to 714 m in the southern native boundary, meaning further reduction of populations within native habitats (Meisner, 1990).
Monterey Cypress (Cupressus macrocarpus) is a species of cypress native to California that thrives in mild and humid climates. The historic distribution of Monterey Cypress forests once spanned the1400 kilometer-long stretch of the California coast from Marin County to Baja California (Graniti, 1998).
The present native distribution is now restricted to a dismal 3.2 kilometer strip on the Monterey Peninsula, where only two relict populations remain (Farjon, 2013). According to the IUCN Red List, the species is classified as threatened and vulnerable with the main forces of the reductions being fire damage and the spread of fungal disease (Farjon, 2013). Cypress canker (Seiridium cardinale), a pathogenic fungus, attacks trees in the cypress family by causing girdling cankers and eventually death of the tree (Graniti, 1998).
While the populations in it’s native territory are dwindling, Monterey Cypress has managed to successfully establish elsewhere. Macrocarpa has been introduced all over the world for use as ornamental trees, windbreaks, and timber (Graniti, 1998). One of the most notable introduction was to New Zealand, where populations flourish. Monterey Cypress was introduced to New Zealand in the 1860s and has since naturalized, finding the climate to be more suitable than that of its native habitat (Wassilleff, 2013). But even in New Zealand, Monterey Cypress has suffered losses due to the spread of cypress canker, causing it to lose popularity for use as timber (Farjon, 2013). Despite the influx of canker in the 1970s, Monterey Cypress still persists throughout rural New Zealand (NZPCN, 2010).
Chinese wattle-necked softshell turtles
The native range of the Chinese wattle-necked softshell turtle (Palea steindachneri) is from the Guangdong region in China down to northern Vietnam. The species is established on Mauritius and the Hawaiian islands Kauai and Oahu, thriving in the warm climate (Ernst & Lovich, 2009). Brought over in the 1800s by Chinese immigrants as a food source, the species has been long established, yet little is known about their present abundance (McKeown & Webb, 1982).
P. steindachneri is an introduced species, and is considered to be potentially invasive, although little is known about their impact and there has been limited research into the ecology and behaviour of the species (Engstrom, 2013; Ernst & Lovich, 2009). Because there is no data available on the growth cycle, population dynamics, or predatory behaviour, it is difficult to estimate the impact their introduction has had and could have on native biota. Current research, led by Dr. Tag Engstrom and Dr. Michael Marchetti from the Center for Ecosystem Research, is investigating the distribution of the softshell turtles on the Hawaiian Islands and their effect on the local ecosystem dynamics (Engstrom, 2013; Radford, 2011).
The wattle-necked softshell turtle is currently listed as endangered on the IUCN Red List due to high demand for turtle products throughout Asia, particularly within China (ATTWG, 2000; Shi et al., 2008).The market for softshell turtle meat for use in traditional Chinese food and medicine is the leading cause of the dramatic population declines. The ongoing capture and trade of the endangered species means that it is unlikely to succeed within its native habitat (Radford, 2011; Shi et al., 2008). The research by Dr. Tag Engstrom and Dr. Michael Marchetti is investigating the invasion/conservation paradox of the softshell turtle and its potential for preservation in its “new” homeland, leading the conversation for the conservation conundrum.
Ex situ conservation
Traditional methods of ex situ conservation involve the translocation or removal of part of a population from its natural habitat to a less threatened location for the preservation of genes or populations. Drawbacks associated with traditional ex situ conservation lie in the inability of the species to thrive within its new habitat due to specific environmental needs. Within the invasion/conservation paradox, the threatened species has already found a more suitable non-native habitat where it has successfully established. Instead of kicking out the intruder, perhaps the populations can be managed closely to allow their persistence.
Within the invasion/conservation paradox, there must be an assessment of risks and benefits, as with many practices of conservation ecology. Do we risk an ecosystem to save one alien species? Or do we eradicate the invader – as most methods of conservation teach – but then risk losing that species entirely? The key is understanding the full impact of the introduced species on the ecosystem it is invading. This is often easier said than done. The complexity of ecosystems and our inability to completely understand all underlying interactions and potential effects makes in difficult for us to anticipate all consequences of introduced species. As with the wattle neck soft-shelled turtle, the species has been naturalized for nearly two centuries and we still don’t know what effects, if any, it has had upon the Hawaiian ecosystem. Yet other species, such as the Monterey Cypress in New Zealand, seem to be perfectly at home within their new habitat without having serious consequences for the native flora and fauna. And then there is the Brook trout, that whilst being damaging to native fishes, is likely to persist in introduced areas due to the demand by anglers. Should this be taken as an opportunity for conservation?
Situations where the non-native species is not considered to be highly detrimental to the native biota create an interesting and new concept of ex situ conservation that could challenge the traditional perception of introduced species. This is a new concept in biology, that requires more questions to be asked, more species to be reevaluated, and more exploration into the many paradoxes that come with the responsibility of conservation.
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Ernst, C., & Lovich, J. (2009). Turtles of the United States and Canada (2nd ed.). Baltimore, Maryland: John Hopkins University Press.
Farjon, A. (2013). Cupressus macrocarpa IUCN 2013 (Version 2013.2 ed.): IUCN Red List of Threatened Species.
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Rieman, B. E., Peterson, J. T., & Myers, D. L. (2006). Have brook trout (Salvelinus fontinalis) displaced bull trout (Salvelinus confluentus) along longitudinal gradients in central Idaho streams? Canadian Journal of Fisheries and Aquatic Sciences, 63(1), 63-78. doi: 10.1139/f05-206
Shi, H., Parham, J., Fan, Z., Hong, M., & Yin, F. (2008). Evidence for the massice scale of turtle farming in China. Oryx, 42(1), 147-150. doi: 10.1017/S0030605308000562
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Biological invasions are happening everywhere. As a main part of global environmental change, they have become a leading cause of the worldwide decline in biodiversity (Mack et al 2000). With such far-reaching ecological effects a lot of effort has been put into finding efficient ways to manage these species. Obviously, the best way to deal with invasive species is to prevent them from invading in the first place (Mack et al 2000). This is, however, not always possible. After establishment, most methods of control are expensive, labor intensive and can have unintended costs. Additionally, for marine environments, where invasions can be much more complex, most of the common methods of prevention and control are not feasible (Atalah et al. 2013a).
As an alternative to the labor-intensive and terrestrial-centric methods, biocontrol has become a popular option. The release of a biological agent, which could be anything from a pathogen to a plant, can be an easier alternative to methods like manual removals or continuous trappings. Howeverthere are many downsides to biocontrol and non-target effects are often associated with these introductions (Thomas & Willis 1998). There is always a danger when introducing a species that they will in turn become a pestand doubly so if it is an exotic species, as is often the case in classic biocontrol (Louda et al. 2003; King & Moody 1982). Avoiding these non-target effects has therefore become a focus for biocontrol efforts. A leading option for this is augmentative biocontrol. Itconsists of bolstering a population that already exists so it can naturally fight off the invaders (Van Lenteren 2012). This would mean significantly decreasing the risk of non-target effects while getting the same amount of control as classical biocontrol.
Sounds perfect, right? Like the panacea for the negative sides of biocontrol?
…Well that’s not quite the case.
A large-scale literature review done by Collier & Van Steenwyk (2004) showed that almost half the time augmentative biocontrol was studied it failed to meet the desired outcome. These failures were based on a multitude of variables including predation, badly chosen environments, and cannibalism within the released species. Additionally, there were times that native populations will almost always be outcompeted just due to ecological circumstances.
Of course this is not to say that augmentative biocontrol is never effective. There are cases where this process can be ecologically and economically the right choice (Atalah & Forrest 2013; van Lenteren 2000; van Lenteren 2012). It has been pointed out that the Collier & Van Steenwyk (2004) review was heavily based on examples in the USA, rather than giving a full view of augmentation applications (van Lenteren 2006). This means that the 50% failure rate is skewed towards a country where augmentative biocontrol is seldom the right choice. In other parts of the world the success rate may be much higher.
Additionally the Collier & van Steenwyk (2004) study is in some part based on poorly selected applications. Usage of augmentative biocontrol on pests which aren’t appropriate for this type of control would skew the results as well. If the review was based on only properly selected applications the success or failure rates may have been extremely different. Below are two such situations where the application of augmentative biocontrol has been demonstrated to have the capacity to control pest species while decreasing non-target effects.
Case 1: Agricultural Pests
Biocontrol has been used heavily for agricultural pests due to its reputation as a safer alternative to chemical herbicides and pesticides (van Lenteren et al. 2003). With agricultural profits being based on the health of the crop species, it seems obvious that non-target effects could have large costs associated with them. Therefore it is necessary to find a way to control pest species without endangering the crop and the insect species needed for successful harvests.
Augmentative biocontrol has been applied to both insect and plant pest species in the agricultural world. Squirting cucumber, a common agricultural plant pest in the Mediterranean Basin, was the focus of one of these studies. The insect Aspongopus viduatus F. was used in an inundative release to define its ability to control the pests. The tests were successful in showing that squirting cucumber was the preferred host of these insects and would therefore be very unlikely to affect similar species used as crops. The prospects for mass rearing and further releases are therefore very good and will be used on a larger scale (Ben-Yakir et al 1996).
In North and South America, and several parts of Asia, Trichogramma spp, are currently being used successfully as control for Lepidopteran pest species, which damage everything from cereal crops to forests (Van Lenteren & Bueno 2003). These parasitic insects work as control by infecting the eggs of the Lepidopteran species through an inundative releases (Smith 1996). In Latin America specifically, this type of augmentative control is used widely as a cost effective control method used to avoid pesticides (Van Lenteren & Bueno 2003).
Case 2: Marine Invasions
Augmentative biocontrol has not been heavily researched in marine environments but it shows a great deal of promise. Invasive seaweeds, kelps and algae have invaded coastal regions around the world and consistently efficient methods of eradication and control have not been found (Anderson 2007). One of the main worldwide invaders is the Asian kelp Undaria pinnatifida. It has been found out of its native range in North and South America, Europe, Australia, and New Zealand (Atalah et al. 2013a; Casas et al. 2008; Thornber et al. 2004). Control of this species is a large concern and the focus of worldwide research. In one of those studies, augmentative biocontrol has been used in conjunction with other methods as a control of Undaria in New Zealand. A large group of sea urchins, Evechinus chloroticus, were transplanted into Breaksea Sound, Fiordland on the South Island. These sea urchins have been found to successfully control Undaria with little to no non-target effects (Atalah et al. 2013a).
Similar types of augmentative biocontrol have also been applied to biofouling of marine structures. Biofouling communities, often found in ports and similarly developed areas, can be composed of many translocated exotic species brought by marine vessels. In these cases natural grazer populations can often not be sustained due to other environmental factors. Due to this fact, augmentative biocontrol has been focused on competitors to exotic pest species (Atalah et al. 2013b). Again New Zealand is leading the way for research into the areas. On the South Island of New Zealand, the population of the sea anemone Anthothoe albocincta was increased through translocations in the Marlborough Sounds. The presence of these sea anemones was found to decrease the amount of problematic pest fouling species (Atalah et al. 2013b). It seems likely that similar augmentative methods, with only a small bit of alteration, could be applied to other marine pests.
So where does augmentative biocontrol stand now?
Even with its limitations, augmentative biocontrol has been successful in the cases discussed above. Additionally it shows promise in other ecosystems like the Australian arid zone where dingoes are being studied as a possible form of feral cat control (Fleming et al. 2012). The methods are, however, applied in a frustratingly small percentage of the world (van Lenteren 2012). This could be due a variety of reasons including a lack of research and perhaps just the fact that many managers are unaware of the idea itself. Or maybe the name “augmentative biocontrol” itself just scared people off the idea. It sounds so overly complicated without really giving you an idea of what it is. Something simpler like augbio or even just augmentation could change things.
As for the limitations found through past research, they are ones that can, in many cases, be controlled for with careful research and site selection. Just like any other form of pest control, You have to identify what species are the best candidates to be controlled. You have to pay careful attention to whether what you’re doing is actually the right thing for what you are trying to control.
Though the uses of augmentative biocontrol are not all encompassing, there is no reason wider applications should not be conducted. There are specific instances where augmentative biocontrol can be extremely effective and these instances are found on a global scale. Success could be seen anywhere that native populations could control an invader at higher numbers or in marine environments where most other control methods for invasive species are not possible. With proper research and well-designed releases augmentative biocontrol could decrease non-target effects of invasive species management worldwide. The success or failure of augmentative biocontrol is entirely up to those people implementing it.
Atalah J, Bennett H, Hopkins G A &. Forrest B M. (2013) Evaluation of the sea anemone Anthothoe albocincta as an augmentative biocontrol agent for biofouling on artificial structures. Biofouling: The Journal of Bioadhesion and Biofilm Research, 29:5, 559-571, DOI: 10.1080/08927014.2013.789503
Atalah J, Hopkins GA, Forrest BM (2013) Augmentative Biocontrol in Natural Marine Habitats: Persistence, Spread and Non-Target Effects of the Sea Urchin Evechinus chloroticus. PLoS ONE. 8(11): e80365. doi:10.1371/journal.pone.0080365
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