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.
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.
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.
ATTWG. (2000). Palea steindachneri IUCN 2013 Red List of Threatened Species. Version 2013.2. Retrieved 1 April, 2014, from http://www.iucnredlist.org/
Colautti, R., & MacIsaac, H. (2004). A neutral terminology to define ‘invasive’ species. Diversity and Distributions, 10, 135-141.
EBTJV. (2013). Eastern Brook Trout: Status and Threats: National Fish Habitat Partnership
Engstrom, T. (2013). The Paradox of Invasive Endangered Species Conservation. Retrieved 1 April, 2014, from http://www.csuchico.edu/cwe/features/tag-engstrom.shtml
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.
Fausch, K. D., & White, R. J. (1981). Competition Between Brook Trout (Salvelinus fontinalis) and Brown Trout (Salmo trutta) for Positions in a Michigan Stream. Canadian Journal of Fisheries and Aquatic Sciences, 38(10), 1220-1227. doi: 10.1139/f81-164
Graniti, A. (1998). Cyrpess canker: a pandemic in progress. Annual Review of Phytopathology, 36, 91-114.
ISSG. (2014). Salvelinus fontinalis (fish) Retrieved 1 April http://www.issg.org/database/species/ecology.asp?si=1226
IUCN. (2013). The ICUN Red List of Threatened Species. Version 2013.2. Retrieved 1 April, 2014, from http://www.iucnredlist.org/
McKeown, S., & Webb, R. (1982). Softshell turtles in Hawaii. Journal of Herpetology, 16(2), 107-111.
Meisner, J. D. (1990). Effect of Climatic Warming on the Southern Margins of the Native Range of Brook Trout, Salvelinus fontinalis. Canadian Journal of Fisheries and Aquatic Sciences, 47(6), 1065-1070. doi: 10.1139/f90-122
NZPCN. (2010). Cupressus macrocarpa. Retrieved 1 April, 2014, from http://www.nzpcn.org.nz/flora_details.aspx?ID=3776
Radford, C. (2011). The endangered wattle-necked softshell turtle (Palea steindachneri) throughout the Hawaiian Islands (Master of Science Thesis), California State University.
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
Warnock, W. G., & Rasmussen, J. B. (2013). Abiotic and biotic factors associated with brook trout invasiveness into bull trout streams of the Canadian Rockies. Canadian Journal of Fisheries and Aquatic Sciences, 70(6), 905-914. doi: 10.1139/cjfas-2012-0387
Wassilleff, M. (2013). Trees in the rural landscape – Macrocarpa and other conifers Te Ara – the Encyclopedia of New Zealand.
“[We must] think beyond our boundaries, beyond ethnic and religious grounds and beyond nations in our global quest for a just world that values and conserves nature.”
– HM Queen Noor of Jordan, opening address at the Vth World Parks Congress, Durban, September 2003
Transboundary Protected Areas (TBPAs), emerging tools for conservation management, hold great potential for the protection and maintenance of biological diversity on the global scale. Originally classified as areas of protected land that cross over a national boundary, the definition of TBPAs has since been expanded to include:
- two or more contiguous protected areas across a national boundary;
- a cluster of protected areas and the intervening land;
- a cluster of separated protected areas without intervening land;
- a transborder area including proposed protected areas; and
- a protected area in one country aided by sympathetic land use over the border
(United Nations Environment Programme)
These cross-boundary protected areas are usually expansive, which can be essential for increasing landscape connectivity and restoring natural habitats. They also allow for greater control over border-specific conservation issues, such as invasive species, illegal trade and poaching, and the reestablishment of large species (UNEP-WCMC).
The concept of TBPAs has gained in popularity over recent years. The number of TBPAs has increased significantly from 59 in 1988 to over 200 in 2007 (Schoon, n.d.). Although the number of TBPAs has improved, our understanding of their success in conserving biodiversity has not. There is a gap in our knowledge, because there is a lack of studies dedicated to monitoring the success of these cross-boundary conservation efforts (Sandwith et al. 2005).How is it that the conservation success of an internationally recognized management tool has been so under-studied?
The Nature of Peace
The very first TBPA was signed into existence in 1924 by Poland and Czechoslovakia under the Krakow Protocol, which “pioneered the concept of international cooperation in establishing border parks.” These protected areas were regarded as a way to reconnect and protect a natural landscape that happened to be divided by an international border, and although international cooperation was vital to their success, the theory of “fostering peace through nature” was not specified as a goal of their formation (Schoon, n.d.). However, over time, the prospect of fostering peace between conflicting nations began to emerge as a key motive for the creation of TBPAs. In 1932, the Glacier-Waterton International Peace Park was established in North America as the first officially declared international peace park. The peace park was dedicated to formally “commemorate the bonds of peace and friendship” between the United States and Canada. The London Convention Relative to the Preservation of Fauna and Flora in their Natural State was signed the following year, which boosted interest in transboundary conservation and called for cross-border cooperation when founding protected areas near political and physical borders (Chester, 2006). Many TBPAs were established in the following years, including the de facto transboundary parks that arose from African national parks following the independence and separation of Rwanda and Democratic Republic of the Congo (formerly Zaire)(Mittermeier et al. 2005).
The array of economic and socio-political benefits that were achieved through cross-boundary cooperation quickly became clear. The term Parks for Peace (or peace parks) started to be used almost interchangeably with Transboundary Protected Areas, and many TBPAs began to be designed to promote goodwill and peace across international boundaries through the conservation of nature (Chester, 2006). There are numerous reports and studies on the socio-political and economic success of cross-boundary partnerships, which is likely to be the reason for their recent popularity (Ali, 2011, & Mittermeier et al. 2005).
Getting Back to Nature
There are two major explanations for why TBPAs are expected to be effective tools for large-scale conservation. First, the commonly large size of TBPAs allows for landscape connectivity across areas that would otherwise be separated by political, social and/or physical boundaries. This connection allows for the uninhibited movement of flora, fauna and ecological processes through the natural landscape (Chassot, n.d.). This can improve the integration of previously separated populations, enable increases in migration, and allow for range adjustment in response to climate change (McCallum et al. n.d.). On a landscape scale, it can also minimize the effects of land use and restore natural habitats. Second, basing management efforts on natural delineations instead of political ones can result in conservation-focused strategies that are more comprehensive and holistic. By pooling the physical, monetary and intellectual resources of two or more countries, the management practices can become more efficient. This can also minimize the impact of border-specific issues, such as invasive species, poaching, and smuggling (McCallum et al. n.d.).
The success of conservation projects is usually measured through follow-up studies on the status of the flora, fauna or ecosystem undergoing mitigation, manipulation or protection. These surveys not only indicate how successful the conservation practices have been, but can also provide insight into how the methods can be altered to improve the effectiveness of the project for this location or a new conservation effort. However, there is a lack of analysis and interpretation of the conservation success of TBPAs – most writing about these protected areas has not been supported by case studies or baseline information (Sandwith et al. 2005). Questions emerge about the ecological effectiveness of TBPAs, the efficiency of TBPAs, and the ability of transboundary conservation initiatives to successfully integrate protection of habitats and biodiversity with the promotion of peace and cooperation (Wolmer, 2004). Is it always necessary to have large, adjoining protected areas across boundaries, or might corridors between existing protected areas be more efficient (Wolmer, 2003)? The many types of TBPAs identified hold very different management and monetary requirements, so without analysis of success it is difficult to prescribe the correct TBPA type with the situation at hand. It is possible that by integrating conservation and socio-economic development programs, one or both of the objectives may suffer (Wolmer, 2004).
Though a shortage of TBPA-specific evidence exists, there are a few relevant projects that provide reason to remain optimistic that transboundary conservation efforts have been, and will continue to be, successful. A transboundary conservation project that integrated the ecosystem management of Kabo-Pokola-Loundoungou forest and the Nouabale-Ndoki National Park in central Africa offers a comprehensive review of the project’s success. A formal management system was implemented to ease communication between the Government of Congo, the Wildlife Conservation Society, and the Congolaise Industrielle des Bois (CIB), a logging company using the forests adjacent to Nouabale-Ndoki National Park.
CIB improved its social and economic standings, while the local communities gained opportunities for employment within the project (Ali, 2011). Several post-implementation studies were undertaken to evaluate the success of the project in terms of fauna distribution. Some of the major findings were that mean species abundance and populations of elephants and gorillas benefited from changes in logging patterns and anti-poaching interventions (Clark et al. 2009, & Stokes, 2010). Although the boundary in this case study was not of political means, the goals and structure of the project are identical to those of international TBPAs. By focusing more efforts on the monitoring of conservation efficacy in TBPAs, there will be a baseline of data with the potential to support further development of cross-boundary protected areas.
Making Peace With Nature
Although it may prove difficult to measure the conservation outcomes of TBPAs, due to an inherent lack of comparable regions and their typically large areas, it is essential to know how effective TBPAs are at meeting their conservation goals. With an increase in political diversity and the diversity of solutions available within conservation projects, there is the potential for major impacts on global biodiversity. The range of TBPA types, from the politically intensive (two or more contiguous protected areas across a national boundary) to the relaxed (a cluster of separated protected areas without intervening land), could provide options that allow for a shift of focus from the outcomes of socio-economic collaborations to the wellbeing of the species and ecosystems involved. The ideal of fostering peace through nature is well researched and is a reasonable mission for TBPAs, as long as we uphold our goal of making peace with nature.
“I know of no political movement, no philosophy, no ideology, which does not agree with the peace parks concept as we see it going into fruition today. It is a concept that can be embraced by all. “
– Nelson Mandela, cofounder of Peace Parks Foundation
Melanie C. Berger is currently undergoing the Masters of Conservation Biology program jointly taught by Victoria University of Wellington in Wellington, New Zealand, and the University of New South Wales in Sydney, New South Wales, Australia. She graduated with an ScB in Biology (with a focus on Ecology and Evolutionary Biology) from Brown University in 2013. She has worked as an Environmental Educator for the NYS Department of Parks, Recreation and Historic Preservation and with the Student Conservation Association and AmeriCorps. She is interested in broadening her knowledge of biogeography, ecology, and conservation biology while making a lasting contribution in these fields. You can learn more about her on her website.
For a Comprehensive View of Transboundary Protected Areas:
For Further Information:
Ali, S. H. (2011). Transboundary Conservation and Peace-building: Lessons from forest biodiversity conservation projects. International Tropical Timber Organization (ITTO) and the United Nations University Institute of Advanced Studies (UNU-IAS).
Chassot, O. (n.d.). Ecological issues – transboundary conservation. TBPA.com. Retrieved April 1, 2014, from http://www.tbpa.net/page.php?ndx=46
Chester, C. (2006). Transboundary protected areas. In Encyclopedia of the Earth (online). Available at: http://www.eoearth.org/article/Transboundary_protected_areas.
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