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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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Headlines and interest group commentary on neonicotinoids


In a paper published this year, the Entomological Society of New Zealand identifies the effects of insecticides and pathogens on honey bees as a “critical issue facing New Zealand entomology” [1]. The paper echoes widespread public and scientific concern regarding the ‘plight of pollinators’ worldwide. Bees have become an emotive topic, subject to increasing politicisation. They provide enormous economic benefits in the form of pollination services to agriculture, as well as New Zealand’s growing apiculture industry. Media, and unfortunately scientists, have muddied the public perception of the problems facing bees through the sensationalising of issues and reporting of unsupported data [2]. The well-publicised phenomenon of colony collapse disorder in North America is a spectre in the public consciousness. Mass bee poisoning events through incorrect insecticide use have made headlines [3]. This increased public awareness has prompted calls from lobbyists [4] to ban insecticide use both overseas and in New Zealand. A new generation of insecticides known as neonicotinoids have become the focus of this debate. Framing the issue in such a narrow, polemic manner is foolish and counterproductive. A caveat of the Entomological Society paper was that further research is needed before measures such as banning insecticides are seriously considered. Such an evidence-based approach to regulation should be followed by New Zealand decision-makers.


Pollinators including honey bees provide pollination services to the world economy worth ~$NZ245 billion per year [5]. In New Zealand it was conservatively estimated in 1987 to be ~$NZ2.2 billion per year [6]. Several of our important crop species including: kiwifruit, onions, pipfruit, berryfruit, avocados, canola and carrot crops are almost solely reliant on honey bee pollination [7]. Besides pollination services, apiculture in New Zealand represents a multimillion dollar industry in its own right. New Zealand honey exports have grown 30% per year in the past decade, reaching NZ$145million in 2013 [8]. New Zealand manuka honey is the world’s most expensive and receives a significant premium over other products. Both pollination services and the apiculture industry provide significant incentives to take a serious approach towards bee health.

honey export volumes, values and prices 2002-2013

Honey export volumes, values and prices: 2002-2013




Honey bee landing on milk thistle. Photo credit: Fir0002

Bee numbers in the developed world appear to be decreasing. In North America, honey bee losses have been particularly acute. There has been a decline from six million hives in the 1940s to 2.3 million today. Mass loss of honey bees from hives has been characterised as colony collapse disorder (CCD). Hypotheses explaining CCD are varied and there is a lack of consensus. The scientific literature points to pathogens, habitat-loss, inadequate bee nutrition, climate change, low genetic diversity, insecticide use and poor bee husbandry practices as causative factors [9]. In Western Europe well-publicised mass bee fatalities due to incorrect insecticide use have made headlines [10] and provided ammunition for commentators in the blogosphere [11].

In New Zealand, anecdotal evidence suggests annual losses of 30-40% of hives by some New Zealand’s apiarists [12]. At present losses are blamed primarily on the effect of the Varroa mite; which increases bee susceptibility to a range of pathogens. However, the New Zealand Bee Keepers Association is also concerned about the use of neonicotinoids on honey bees [13].


In recent years a series of high-impact studies showing deleterious effects of neonicotinoid insecticides on bees have been published. Some of these effects include: significant loss of queens [14], interference with foragers’ ability to navigate back to the hive [15] and synergistic deleterious effects of insecticides on bee colonies at low dosages [16].

Studies such as these, as well as media coverage and public pressure have prompted intergovernmental bodies such as the European Food Safety Authority (EFSA) to become involved. As of December 2013 the EU, acting on EFSA recommendations, has restricted field uses of three neonicotinoid insecticides. Unsurprisingly these actions elicited a strong response from neonicotinoid manufacturers. Syngenta called the risk assessment conducted by the EFSA, hurried and inadequate”. Bayer stated it does not believe the EFSA’s reports “alter the quality and validity” of previous risk assessments by the EU.

Domestic agencies have also waded in. In the United Kingdom, the Department for Environment, Food and Rural Affairs has concluded that there is a “growing body of evidence” that neonicotinoids do not exert an effect under conditions where bees can forage naturally. This suggests that studies in the lab that have linked neonicotinoids to an impact on bee health at less-than-lethal doses “did not replicate realistic conditions, but extreme scenarios”.

Closer to home, a report was released by the Australian Pesticides and Veterinary Medicines Authority (APVMA) in February of this year on neonicotinoids and the health of honey bees in Australia. The APVMA concluded the introduction of neonicotinoids has led to an overall reduction in the risks to the agricultural environment from the application of insecticides. The reasons given were extensive, but can be reduced to the following:

  1. Previous generation insecticides are significantly more toxic to mammals;
  2. The physicochemical properties of some neonicotinoids means they are used to coat crop seeds, which protects the seed and plant while they are growing. Consequently this means farmers apply fewer insecticides to the environment;
  3. Alternative application methods have a higher potential to introduce insecticides into the environment; and
  4. In Australia honey bee populations are not in decline despite extensive use of neonicotinoid insecticides.

However, potential risks of neonicotinoids are also identified. In particular, the physicochemical properties of neonicotinoids that allow them to translocate from the seed to the growing plant mean they have high stability in plant tissue and soil. Particularly worrying is that neonicotinoids may contaminate nectar and pollen which is a main food source for bees [17]. Neonicotinoids therefore could represent a larger environmental threat than other less mobile, less persistent insecticides.

Given the work occurring internationally and the importance of honey bees to New Zealand, the question arises: What action should we take?


Popular neonicotinoid products. You might even recognise some from your garden at home.


New Zealand’s environmental regulatory regime is well-developed. The legislature does a relatively good job of incorporating science into policy, regulation and operational decision-making. Such an evidence-based approach to regulation should apply to the debate regarding neonicotinoid use and bees. However, how ‘precautionary’ should that approach be?

Precautionary Approach? Other solutions

The precautionary approach had its genesis in international environmental law and dictates that lack of scientific evidence of harm is no reason not to legislate against harm occurring. The principle has long been accepted to apply in the areas of food safety and plant and animal protection [18]. Further, it is enshrined in the Convention on Biological Diversity 1992, to which New Zealand is a party. It is also a fundamental purpose of the HSNO Act [19]. Should the precautionary approach apply to neonicotinoid use in New Zealand? In resource management case law, it has been made clear that the weight given to the precautionary approach depends on circumstances involved [20].


As this article has demonstrated, many published studies have convincingly described sublethal effects of neonicotinoids on bees in laboratory conditions. However, to this author’s knowledge, no adverse effects to bee colonies has ever been observed in field studies at field-realistic exposures [21]. Additionally, an enormous body of literature attributes decline in bee health not solely to insecticides, but to pathogens, habitat-loss, inadequate bee nutrition, climate change, low genetic diversity and poor bee husbandry practices.

Although the precautionary approach is potentially applicable in this scenario it is not appropriate to apply in favour of banning neonicotinoids; the circumstances do not recommend it. Instead, a twofold approach should be pursued:

  1. First, there is clear evidence Varroa decimates New Zealand honey bee colonies [22]. Research should be undertaken both to understand bee pathogens’ impacts and combat the effect they have on New Zealand bee populations.
  2. Second, if any measures banning insecticides are to be considered, extensive field research and risk-analyses are required. A first step would involve testing what impacts sub-lethal effects observed in the lab have in the natural environment.



Honey bee carrying pollen. Photo credit: Muhammed Mahdi Karim.

Misinformation and sensationalist media coverage has heightened public awareness of the challenges facing bees. The
‘plight of pollinators’ is not a minor issue, nor is it being treated as such.  Intergovernmental and domestic agencies around the world are approaching the problem with due seriousness. The effects of banning neonicotinoids are at present murky. Indeed, other insecticides more harmful to bees and mammals could be reintroduced. Causative factors of bee decline in the developed world are not yet fully elucidated. However diseases, not insecticides, are likely the primary cause of the losses. Framing the debate to focus on neonicotinoids is an irresponsible approach by commentators that delivers political capital whilst attenuating efforts to deal with the problem holistically.  The potential application of a precautionary approach to neonicotinoids should be rejected in New Zealand given the possible negative trade-offs. Overly parochial views of lobbyists should be dismissed by decision-makers in favour of evidence-based regulation. If evidence suggests change in the risk management of any substance is required, then the regulatory system should respond accordingly.


[1] Lester PJ, Brown SDJ, Edwards ED, Holwell GI, Pawson SM, Ward DF & Watts CH (2014) Critical issues facing New Zealand entomology. New Zealand Entomologist 37: 1-13.

[2] Maini S, Medrzycki P & Porrini C (2010) The puzzle of honey bee losses: A brief review. Bulletin of Insectology 63: 153-160.

[3] http://www.theguardian.com/environment/2008/may/23/wildlife.endangeredspecies

[4] see https://www.greens.org.nz/bees + http://tvnz.co.nz/world-news/nz-should-follow-eu-ban-bee-harming-pesticides-greens-5422307 + http://organicnz.org.nz/node/690

[5] Gallia N, Salles JM, Settele J et al. (2009) Economic valuation of the vulnerability of world agriculture confronted with pollinator decline. Ecological Economics 68: 810-821.

[6] Matheson AG & Schrader M (1987) The value of honey bees to New Zealand’s primary production. Ministry of Agriculture and Fisheries Report.

[7] Ministry for Primary Industries 2013 apiculture monitoring programme, p 3.

[8] Ministry for Primary Industries 2013 apiculture monitoring programme, p 1.

[9] Bromenshenk JJ, Henderson CB, Wick CH, Stanford MF, Zulich AW, Jabbour RF et al. (2010) Iridovirus and microsporidian linked to honey bee colony decline. PLoS One: e13181; Evans JD & Schwarz RS (2011) Bees brought to their knees: microbes affecting honey bee health. Trends in Microbiology 19: 614-620; Henry M, Beguin M, Requier F, Rollin O, Odoux JF, Aupinel P et al. (2012) A common pesticide decreases foraging success and survival in honey bees. Science 336: 348-350; Cresswell JE (2011) A meta-analysis of experiments testing the effects of neonicotinoid insecticide (Imidacloprid) on honey bees. Ecotoxicology 20: 149-157; Winfree R, Bartomeus I & Cariveau DP (2011) Native Pollinators in Anthropogenic Habitats. Annual Review of Ecology Evolution and Systematics 42: 1-22.

[10] http://www.theguardian.com/environment/2008/may/23/wildlife.endangeredspecies

[11] http://ecowatch.com/2013/04/19/greenpeace-syngenta-pesticides-kill-bees/ http://www.theguardian.com/environment/georgemonbiot/2013/aug/05/neonicotinoids-ddt-pesticides-nature

[12] Lester PJ, Brown SDJ, Edwards ED, Holwell GI, Pawson SM, Ward DF & Watts CH (2014) Critical issues facing New Zealand entomology. New Zealand Entomologist 37: 1-13.

[13] Foster B (2013) Our science challenge. New Zealand Bee Keeper 21: 4-6.

[14] Whitehorn PR, O’Connor S, Wackers FL & Goulson D (2012) Neonicotinoid Pesticide Reduces Bumble Bee Colony Growth and Queen Production. Science 336: 351–352.

[15] Henry M, Beguin M, Requier F, Rollin O, Odoux JF, Aupinel P et al. (2012) A common pesticide decreases foraging success and survival in honey bees. Science 336: 348-350.

[16] Gill RJ, Ramos-Rodriguez O & Raine NE (2012) Combined pesticide exposure severely affects individual – and colony – level traits in bees. Nature 491: 105-109.

[17] Rortais A, Arnold G, Halm MP & Touffet-Briens F (2005) Modes of honeybees exposure to systemic insecticides: estimated amounts of contaminated pollen and nectar consumed by different categories of bees. Apidologie 36: 71-83.

[18] (New Zealand v Japan; Australia v Japan) Provisional Measures ITLOS Case Nos 3 & 4, 27 August 1999; EC Measures Concerning Meat and Meat Products (Hormones) WTO DOC WT/DS26/AB/R (1998) (Report of Appellate Body); New Zealand Pork Industry Board v Director-General of Ministry for Primary Industries [2013] NZSC 154.

[19] Section 7, Hazardous Substances and New Organisms Act 1996.

[20] McIntyre v Christchurch City Council (1996) NZRMA 289 (PT).

[21] Blacquiere T, Smagghe G, van Gestel CAM & Mommaerts V (2012) Neonicotinoids in bees: a review on concentrations, side-effects and risk assessment. Ecotoxicology 21: 973-992.

[22] Howlett BG & Donovan BJ (2010) A review of New Zealand’s deliberately introduced bee fauna: current status and potential impacts. New Zealand Entomologist 33: 92-101.

Dredging on the Great Barrier Reef: Can resource use and conservation co-exist?- Malindi Gammon


The Great Barrier Reef is a multi-use marine park in which conservation is combined with resource use such as commercial fishing and tourism. Plans to expand an existing coal port within the region have recently been approved. The government plans to coincide such resource use with conservation, ensuring any adverse environmental effects are avoided or mitigated.        However, several factors surrounding the approval of Abbot Port expansion are in conflict with core conservation values: 1.) The reef is already in a state of jeopardized health as a direct result of anthropogenic stressors, 2.) The expansion will lead to some ongoing effects which can’t be mitigated and 3.) The expansion is to allow for capital investment in coal energy, an unsustainable energy source. In the instance of Abbot Port expansion, resource use and conservation can not co-exist.


The Great Barrier Reef (GBR) is a world Heritage site acknowledged for its outstanding global significance to biodiversity. Covering over 348,000 square kilometres, the Great Barrier Reef (GBR) encompasses approximately 3000 reefs and represents 10% of the world’s reefs (GBRMPA, 2009). The GBR is home to many unique and endangered species, including over 1500 species of fish, 350 species of hard coral, 6 of the world’s 7 species of marine turtle, the dugong and more than 30 species of whale and dolphin (Wachenfeld et al, 2007). The GBR is a delicate ecosystem, existing on a fine balance between the marine environment and the interactions of all species living there. This balance creates a spectacular array of colour and life forms, upon which Australia has built its international identity (Figure 1).           The value of the GBR isn’t restrained to its ecological value, it also holds immense monetary value and resources. The GBR is a multiple use marine park which is open to sustainable resource use and supports a commercial marine tourism and fishing industry. An estimated 6 million tourists visit the GBR annually, contributing $6.1 billion dollars to the Australian tourism industry (Wachenfeld et al, 2007). 


Figure 1: This photo illustrates the amount of life which exists even within a small sub-section of the vast reef. Small reef fish are taking refuge in stony coral whilst larger fish school above. Retrieved from: http://ngm.nationalgeographic.com/2011/05/great-barrier-reef/doubilet-photography.

On 10 December 2013 the Australian government approved a dredging programme for proposed terminals at the Port of Abbot, a port which is located within the Great Barrier Reef Marine Park and used for the exportation of coal (MFE, 2013a). This proposal included the dredging of up to 3 million cubic metres of spoil (sand, silt and clay off the seafloor) and the disposal of this 24km off-shore from the dredging site (MFE, 2013b). The decision to approve this project has not been taken lightly, and the conclusion was reached under the agreement of 47 strict environmental conditions, including:  150% net benefit requirement for water quality, approximate monetary contribution to projects supporting reef health of $89 million and measures for protection of marine species and communities (GBRMPA, 2009). Despite these stringent conditions, the approval has been met with a fierce public debate (Fig. 2).


Figure 2: A crowd protests the approval of dredging at Point Abbot and dumping of spoil within the Great Barrier Reef Marine Park. Photo: The Cairns Post, February 04 2014.

Prior to approval, the proposal attracted 228 submissions in opposition (The Department of State Development, Infrastructure and Planning, 2013). These submissions, were from individuals, consultancies and both nongovernmental and governmental organisations and cited adverse environmental impacts as a primary cause for concern (The Department of State Development, Infrastructure and Planning, 2013). The GBR Marine Park Authority aims to balance economic development with environmental protection, stewardship and conservation. I pose the question of whether resource use and conservation can co-exist in the context of dredging on the GBR, as it has done in the past with commercial fishing and tourism.


Two sides to the debate:

The Australian economy stands to benefit substantially from the proposed dredging and expansion of Abbot Port. Expected outcomes include: an additional $660 million of revenue per year, $123 million household income per year and 2,300 full-time jobs (Ports Corporation of Queensland, 2008). In addition to the massive economic gain, stringent environmental conditions have been set in place to limit negative impacts.  In his press release the Minister for the Environment Hon. Greg Hunt, stated that the approval was granted under strict environmental conditions to avoid, remedy and mitigate any adverse environmental impacts (MFE, 2013). In particular, the condition to ensure net water quality state of 150% its current state, will greatly improve the quality of water within GBR to a level beyond what it is currently. Terrestrial run-off of polluted water and its effect on water quality has a significant adverse effect on many species within the GBR (Schaffelke et al, 2005). The outcome of this is to ensure a positive gain, both environmentally and economically, from this proposed development.               A 2009 outlook report for the GBR cited climate change (increasing sea temperature, ocean acidification and rising sea level), catchment runoff, sedimentation and coastal development as the greatest threats to the health of the GBR (GBRMPA, 2009). The effects of dredging and mining on the reef were not cited as a major source of degradation, a fact dutifully mentioned by those in support of the Port of Abbot expansion. The minimum $89 million contribution to support projects aimed at reef health could make steps towards researching and mitigating these major causes of decline.

When considering the effect that the Abbot Port expansion may have on the GBR, we must first consider the ecosystems current state of health and it’s resilience to environmental stressors. This ecosystem is already exposed to a multitude of anthropogenic stressors, both direct and indirect:  Ocean acidification (De’ath et al, 2009), large scale-bleaching events (Berkelmans et al, 1999) and rising sea water temperatures (Hoegh-Guldberg et al, 2007). All of these factors have led to a loss of over half the initial coral cover since 1985 (De’ath et al, 2012) (Fig. 3)



Figure 3: Box plots of the percentiles (25%, 50% and 75%) of coral cover distributions within each year at the GBR. A significant decline since 1985 is evident of 28% coral cover to 13.8% coral cover (De’ath et al, 2012).

Not only is the reef in a state of jeopardised health, but, despite attempts to alleviated and mitigate any adverse effects during dredging and construction, the reef is still likely to be negatively impacted. Some species of coral are very sensitive to sedimentation. Sedimentation will result during dredging, and dumping due to a large expanse of fine particles being released into the water column. Certain coral species can be extremely sensitive to the sedimentation of disrupted spoil during dredging (Erftemeijer et al, 2012). Due to the varying tolerance of many coral species, some corals will be extremely sensitive to sedimentation, whilst those with a greater tolerance will likely thrive in the absence of less tolerant species. This could cause a shift in the community composition towards a dominance in sedimentation resilient species (Sofonia et al, 2008).  Such a shift in coral community dominance could have a bottom-up effect on the wider community and those species which rely on particular coral as a form of nutrients and protection. Although such an extreme effect is unlikely, affected corals are in a prior state of stress and we do not know how close to their maximum tolerance levels they may be.                   The dredging is part of an expansion plan for Abott Port, a port used in the exportation of coal. If this expansion were to occur, more coal transporting vessels will pass through and within the vicinity of the reef. Juvenile reef fish use sound created by the reef to locate habitat and settle (Radford et al, 2011). Sound created by coal transporters would likely “drown-out” the sound of the reef and limit fish larvae’s ability to locate and settle on the reef (Holles et al, 2013). An increase in shipping traffic would also increase the chance of invasive species introduction via ship ballast waters (Lavoie et al, 1999). Both these factors are outcomes which need to be considered carefully as there impacts go well beyond the initial stages of development. Not only will dredging cause an initial and ongoing disturbance to an ecosystem which is already under considerable anthropogenic stress, but the dredging is being undertaken to expand investment into an unsustainable resource, coal. Consideration has to be given to the affect such an investment will have on current attempts to mitigate climate change and shift toward renewable sources of energy.


The approval of Abbot Port expansion was not met without careful environmental consideration and stringent conditions aimed to protect the environment and mitigate any negative impacts. If the GBR where existing in isolation, sealed off from current environmental issues plaguing the world, then these conditions would likely suffice. However, this is not the case. The GBR is already subjected to immense anthropogenic stressors as evident by mass bleaching (Berkelmans et al, 1999), reduction in coral cover (De’ath et al, 2009) and a general decline in reef health. All these factors have reduced the resilience of the reef, and any further impact should be avoided.     Several outcomes of the Abbot Port expansion are at conflict with conservation values. Despite best-practise attempts to reduce any sedimentation effects, some is likely to occur. Many coral species are sensitive to sedimentation (Erftemeijer et al, 2012) and due to the current state of reef health (De’ath et al, 2012) we don’t know how resilient these species may be. Some ongoing effects can not be remedied, especially the effect additional boat noise may have on larvae settlement (Holles et al, 2013). Finally, the dredging is to allow for the expansion of a port which is used for the exportation of coal. Coal, being a fossil fuel, is an unsustainable energy source. The combustion of coal contributes greatly to global pollution and carbon dioxide level, both of which have put the GBR under considerable stress. Can resource use and conservation co-exist in the context of dredging on the Great Barrier Reef? My answer is No.


Australian Government: Great Barrier Reef Marine Park Authority (GBRMPA). (2009). Great Barrier Reef Outlook Report 2009: In brief. Great Barrier Reef Marine Park Authority: Queensland, Australia.

Australian Government: Ministry for the Environment (MFE). (2013a). Abbot Point and Curtis Island projects approved- New safeguards to protect the long-term future of the Great Barrier Reef. [Press release]. Retrieved from: http://www.environment.gov.au/minister/hunt/2013/pubs/mr20131210.pdf. Retrieved on: 02.04.2014.

Australian Government: Ministry for the Environment (MFE). (2013b) Abbot Point and Port of Gladstone Projects Summary. [Press release]. Retrieved from: http://www.environment.gov.au/minister/hunt/2013/pubs/abbot-point-projects.pdf. Retrieved on: 02.04.2014.

Australian Government: Great Barrier Reef Marine Park Authority (GBRMPA). (2014). Permit G14/34897.1. Retrieved from: http://www.gbrmpa.gov.au/__data/assets/pdf_file/0019/123166/G34897.1-signed.pdf. Retrieved on: 02.04.2014.

Berkelmans, R., & Oliver, J. K. (1999). Large-scale bleaching of corals on the Great Barrier Reef. Coral reefs18(1), 55-60.

The Department of State Development, Infrastructure and Planning (2013). Great Barrier Reef Ports Strategy Consultation Report Version 1.1, summary of consultation responses. Australia: Queensland.

De’ath, G., Lough, J. M., & Fabricius, K. E. (2009). Declining coral calcification on the Great Barrier Reef. Science323(5910), 116-119.

De’ath, G., Fabricius, K. E., Sweatman, H., & Puotinen, M. (2012). The 27–year decline of coral cover on the Great Barrier Reef and its causes.Proceedings of the National Academy of Sciences109(44), 17995-17999.

Erftemeijer, P. L., Riegl, B., Hoeksema, B. W., & Todd, P. A. (2012). Environmental impacts of dredging and other sediment disturbances on corals: a review. Marine Pollution Bulletin64(9), 1737-1765.

Hoegh-Guldberg, O., Mumby, P. J., Hooten, A. J., Steneck, R. S., Greenfield, P., Gomez, E., & Hatziolos, M. E. (2007). Coral reefs under rapid climate change and ocean acidification. science318(5857), 1737-1742.

Holles S., Simpson S. & Radford A. (2013). Boat noise disrupts orientation behaviour in coral reef fish. Marine Ecology Progress Series 485: 295-3000.

Lavoie, D. M., Smith, L. D., & Ruiz, G. M. (1999). The potential for intracoastal transfer of non-indigenous species in the ballast water of ships. Estuarine, Coastal and Shelf Science, 48(5), 551-564.

Ports Corporation of Queensland (2008). Report for Abbot Point Coal Terminal X110 Expansion. Australia: Queensland. [Retrieved from: http://d301432.u111.fasthit.net/files/Submitted_EPBC/Port/Attachments/Attachment%20No.1%20PCQ%20Documents/IAS_19092008%5B1%5D.pdf].

Radford, C. A., Stanley, J. A., Simpson, S. D., & Jeffs, A. G. (2011). Juvenile coral reef fish use sound to locate habitats. Coral Reefs, 30(2), 295-305. 

Schaffelke, B., Mellors, J., & Duke, N. C. (2005). Water quality in the Great Barrier Reef region: responses of mangrove, seagrass and macroalgal communities. Marine Pollution Bulletin, 51(1), 279-296.

Sofonia, J. J., & Anthony, K. (2008). High-sediment tolerance in the reef coral< i> Turbinaria mesenterina</i> from the inner Great Barrier Reef lagoon (Australia). Estuarine, Coastal and Shelf Science78(4), 748-752.

Wachenfeld, D., Johnson, J., Skeat, A., Kenchington, R., Marshall, P., & Innes, J. (2007). Introduction to the Great Barrier Reef and climate change. Climate change and the Great Barrier Reef: a vulnerability assessment, 1-13.

Breaking out of the Fortress: A Biocultural Approach to Conservation

– Te Taiawatea Moko-Mead

“Perhaps the greatest challenge of all is to change the way we think about protected areas. In the past they have been seen as islands of protection in an ocean of destruction. We need to learn to look on them as the building blocks of biodiversity in an ocean of sustainable human development, with their benefits extending far beyond their physical boundaries”

Achim Steiner, New Scientist 18 October 2003. p21

Ngāti Awa signing of the deed of settlement, a gateway for indigenous management approaches

Ngāti Awa signing of the deed of settlement, a gateway for indigenous management approaches. Photo contribution: Aroha Mead


Biodiversity is declining worldwide, with current extinction levels rising rapidly (Gorenflo et al., 2012). This decline is co-occurring with the knowledge systems that are interconnected with and have long supported biodiversity (Gorenflo et al., 2012; Maffi & Woodley, 2012; Stephenson et al., 2014). However, when confronting biodiversity conservation, management approaches focus little on addressing both human and biological needs (Gavin et al., 2014). These approaches remain topics of constant debate, and communicates the division between people-orientated and fortress conservation approaches. Fortress conservation seeks to preserve the environment through forceful exclusion of people, including indigenous and local communities whom have relied on the environment for their livelihoods (Brockington, 2002; Wilshusen et al., 2002).

I aim to critiqe the fortress conservation approach mainly because it; silences the rights of indigenous and local communities to their traditional lands and resources (Fabricius et al., 2001; Goetze, 2005; Wadley, 2002), resources are too complex to be governed by a single agency (Berkes, 2009) and hands on resource use and long-term commitment to sustaining resources can produce greater conservation outcomes (Stephenson et al., 2014). An emerging field of biocultural conservation aims to address the injustices encountered by this approach and instead links biophysical and socio-cultural components in a social-ecological system (Gavin et al., 2014), where a more hands on approach to conservation management is used. An example will be examined throughout, focusing on biocultural conservation through work analysed by (Stephenson et al., 2014) in a co-management agreement between indigenous Māori and a Crown agency in New Zealand.

Evaluating the fortress paradigm

So why is this approach used? Wilshusen et al., 2002 has outlined five arguments which form the foundation of the fortress conservation approach, these arguments include that;

  1. Protected areas require strict protection.
  2. Protecting biodiversity is a moral imperative.
  3. Conservation linked to development does not protect biodiversity.
  4. Harmonious, ecologically friendly local communities are myths.
  5. Emergency situations require extreme measures.

Although these arguments have a substantial amount of literature and significant backing behind them, the foundation of these arguments are extremely flawed. Several assumptions are made and they overlook the fact that protection alters social and political landscapes, they mask the fact that ‘conservation for the common good’ refers to the special interests of the elite, it assumes the idea that the government serves the common good of their citizens, it also assumes that local institutions cannot adapt to social change along with many other assumptions (Wilshusen et al., 2002). This approach is supposedly in the good name of conservation but involves methods which are socially unjust and have proven to produce less efficient outcomes, in comparison to bottom-up approaches to conservation (Ostrum, 1990; Wilshusen et al., 2002).

..So what are the alternatives?

The emerging field of biocultural conservation

Biocultural conservation aims to achieve goals which supports human and biological needs, and bridge the gap between scientific biodiversity conservation and local and indigenous values of biodiversity (Gavin et al., 2014; Stephenson et al., 2014). This has recently been defined as;

‘Conservation actions made in the service of sustaining the biophysical and social-cultural components of dynamic, interacting and inter-dependant social-ecological systems’ – (Gavin et al., 2014).

However, can biocultural conservation provide the right framework to address the injustices created by fortress conservation?

(Gavin et al., 2014) has defined ten principles to guide successful biocultural conservation initiatives, some of these principles include;

  1. Acknowledging that conservation can have multiple objectives and stakeholders.
  2. Recognizing that culture and language are dynamic, and that this dynamism shapes resource use and conservation.
  3. Respect that different worldviews shape human-environment interactions, and incorporate these differences into conservation planning.
  4. Prioritize the importance of partnership and relationship building for conservation outcomes.

The ironic thing about biocultural conservation is that, it is already an integral part of resource use and management of indigenous people. As their practices reflect a long history of co-evolving and interdependent social-ecological systems (Gavin et al., 2014; Stephenson et al., 2014). Fortunately, settlements of indigenous rights issues has provided a platform for indigenous management approaches to be applied (Coates, 2009; Stephenson et al., 2014).

Case study: biocultural conservation of fisheries in New Zealand  

In New Zealand, Māori (the indigenous peoples) have sought settlements from the Crown to make up for past grievances and confiscation of land and resources in early Colonisation (Linkhorn 2011). These settlements are usually bound with co-management arrangements between Maori and the Crown for protected areas. Fisheries in New Zealand will be used as an example to illustrate this.

There are two main approaches to management of marine reserves; firstly most fish stock in EEZ (Exclusive economic zone) are privatised or allocated under QMS (Quota management system); and secondly the principal approach to conservation is to set marine reserves aside where fishing is prohibited (MPI, 2009; Stephenson et al., 2014). This approach of all or nothing does not match well with the customary approach. According to local iwi (Māori tribal groups) it is neither resulting in more sustainable local fisheries (Stephenson et al., 2014).

Fisheries is a traditional source of economic and cultural wealth for Maori. Although, Māori were not legally able to be involved in fisheries through extensive negotiations and Treaty of Waitangi settlements the crown were forced to create regulations to allow traditional fisheries to be managed by Māori (Coates, 2009). Three management tools were able to be established because of this, these are co-management arrangements between iwi and crown agencies, in this case Ministry for Primary Industries.

These management tools demonstrate a biocultural approach to conservation in New Zealand and include; Taiāpure, Mātaitai and Rāhui (s186 temporary closures).

These reserves are defined by the Ministry for Primary (MPI, 2009) as;

Mātaitai: A gazetted area where tangata whenua establish a reserve on a traditional fishing ground for the purpose of recognising and providing for customary management practices and food gathering. Commercial fishers may not fish in a mätaitai reserve, however recreational fishers can.

Taiāpure: established in an area that has customarily been of special significance to an iwi or hāpu as a source of food or for spiritual or cultural reasons. All fishing (including commercial fishing) can continue in a taiāpure. This tool offers a way for Tāngata Whenua to become involved in the management of both commercial and non-commercial fishing in their area.

Rāhui (s186 temporary closures): Section 186 A (North & Chatham Island) & B (South Island) of the Fisheries Act 1996 allows the Minister of Fisheries or the Chief Executive of the Ministry of Fisheries to temporarily close an area to fishing. The specific purpose is to provide for the use and management practices of Tāngata Whenua in the exercise of their customary rights this approach is otherwise known as a rāhui.

These methods of conservation have proven to be a success, where an example of a Taiapure reserve has had notable success in the Bay of Plenty. A Taiapure committee is set up composed of tribal representatives and local recreational and commercial fishers whom regularly monitor the area. They have also had considerable success in effecting policy changes to support local customary practices(Stephenson et al., 2014).

Area of government established coastal marine reserves compared to the indigenous co-managed mātaitai and taiāpure (Stephenson et al., 2014)

Figure 1. Area of government established coastal marine reserves compared to the indigenous co-managed mātaitai and taiāpure (Stephenson et al., 2014)

It is also noted that since 1994, the number of taiapure and maitaitai reserves have risen exponentially (Figure 1). Not including the two large offshore reserves around the Kermadec and Auckland Islands, the area protected as mātaitai and taiāpure is about twice that of coastal marine reserves (Stephenson et al., 2014).

The management tools offer opportunities for Māori to be more actively engaged in the management of their local fisheries, while also producing efficient conservation outcomes (Stephenson et al., 2014).

The barriers and constraints of biocultural conservation

We still have a long way to go…

(Mulrennan & Scott, 2005) recites a co-management agreement in Australia between Torres Straight Islanders and the government. Where a consultative and advisory structure was established under the Treaty. This involved Islander, central government and industry representatives. The Torres Strait Protected Zone Joint Authority (TSPZJA) represented the top-level fisheries decision-making committee. However through this co-management agreement, indigenous institutions of land and sea tenure, resource management and environmental knowledge had little effect on the management institutions, processes and decisions, where the role of indigenous was simply advisory.

Although the future of biocultural conservation seems promising, there are still many barriers which we need to overcome. These barriers include; significant sharing of power across levels, funding, and a struggle to adjust to the dynamic nature of social-ecological systems(Gavin et al., 2014). Where we lay an emphasis on meaningful power-sharing, as dynamic and polycentric. Taiepa et al., 1997 articuates that the kereru (Figure 2.) shows the potential for developing partnerships and ‘bridge building between the different world views.  Where returning the life essence whakahokia te mauri , self determination rangatiratanga,  guardianship kaitiakitanga and customary law tikanga are just as important in co-management as ecology and ecosystem health.

The kereru, a symbol of meaningful co-management in New Zealand (Taiepa et al., 1997)

Figure 2. The kereru, a symbol of meaningful co-management in New Zealand (Taiepa et al., 1997)


As biocultural conservation is an emerging field, the success of these areas in achieving goals of supporting Indigenous peoples and their rights, and aiding sustainabilty and increasing the species and environments within their territories, is yet to be decided (Gavin et al., 2014; Stephenson et al., 2014). But the opportunities for communities involved is significant; greater participation of local people, greater recognition of knowledge systems intertwined with conservation and a greater role in the decision-making. Where, locally driven approaches may be the best hope for long-term conservation(Stephenson et al., 2014). This enables a bottom-up approach to conservation where communities and indigenous peoples are involved in hands on sustainable resource management.

In conclusion, conservation should aim to recognise the dense web of social and political processes along with conservation to provide socially just methods of conservation, where restoring human interaction should also be prioritised (Maffi & Woodley, 2012). We should not build a fortress around the environment and simply turn a blind eye to the implications. Instead we should focus on breaking down the fortress and interacting with the communities, indigenous peoples and the environment of which we have co-evolved with. We need to get past power and control in conservation and remember that people and nature have successfully co-existed for generations. Conservation does not need to be a trade-off.


Aroha Mead: Chair of IUCN Commission on Environmental, Economic and Social Policy, IUCN Councillor, Māori Business Director (Victoria University of Wellington)

Kim Turrell: Honours candidate Victoria University of Wellington


Berkes, F. (2009). Evolution of co-management: Role of knowledge generation, bridging organizations and social learning. Journal of Environmental Management, 90(5), 1692-1702. doi: http://dx.doi.org/10.1016/j.jenvman.2008.12.001

Brockington, D. (2002). Fortress conservation: the preservation of the Mkomazi Game Reserve, Tanzania: Indiana University Press.

Coates, N. (2009). Joint-management agreements in New Zealand: simply empty promises. Journal of South Pacific Law, 13(1), 32-39.

Fabricius, C., Koch, E., & Magome, H. (2001). Towards strengthening collaborative ecosystem management: lessons from environmental conflict and political change in southern Africa. Journal of the Royal Society of New Zealand, 31(4), 831-844.

Gavin, M., McCarter, J., Mead, A., Berkes, F., Stepp, J., Peterson, D., & Tang, R. (2014). Defining Biocultural Approaches to Conservation. manuscript to be published.

Goetze, T. C. (2005). Empowered Co-Management: Towards Power-Sharing and Indigenous Rights in Clayoquot Sound, BC. Anthropologica, 47(2), 247-265. doi: 10.2307/25606239

Gorenflo, L. J., Romaine, S., Mittermeier, R. A., & Walker-Painemilla, K. (2012). Co-occurrence of linguistic and biological diversity in biodiversity hotspots and high biodiversity wilderness areas. Proceedings of the National Academy of Sciences, 109(21), 8032-8037.

Maffi, L., & Woodley, E. (2012). Biocultural diversity conservation: a global sourcebook: Routledge.

MPI. (2009). Customary Management from http://www.fish.govt.nz/en-nz/Maori/Management/default.htm

Mulrennan, M. E., & Scott, C. H. (2005). Co-management – An Attainable Partnership? Two Cases from James Bay, Northern Quebec and Torres Strait, Northern Queensland. Anthropologica, 47(2), 197-213.

Ostrum, E. (1990). Governing the commons. Cambridge University-Press, Cambridge.

Stephenson, J., Berkes, F., Turner, N., & Dick, J. (2014). Biocultural conservation of marine ecosystems: Examples from New Zealand and Canada Indian Journal of Traditional Knowledge, 13(2), 257-265.

Taiepa, T., Lyver, P., Horsley, P., Davis, J., Bragg, M., & Moller, H. (1997). Co-management of New-Zealand’s conservation estate by Maori and Pakeha: a review. Environmental Conservation, 24(3), 236-250.

Wadley, R. L. (2002). The history of displacement and forced settlement in West Kalimantan, Indonesia – implications for co-managing Danau Sentarum Wildlife Reserve.

Wilshusen, P. R., Brechin, S. R., Fortwangler, C. L., & West, P. C. (2002). Reinventing a square wheel: Critique of a resurgent” protection paradigm” in international biodiversity conservation. Society &Natural Resources, 15(1), 17-40.

About the author:

IMG_3260Kia ora koutou, my name is Te Taiawatea Moko-Mead.  I am from Aotearoa, New Zealand, my iwi (tribal groups) are; Ngāti Porou, Ngāti Awa and Tainui. I am a student at Victoria University of Wellington, studying a Master of Conservation Biology degree. I have previously studied a BSc in Marine Biology, minors in Environmental Studies and Statistics. My areas of interests are in; co-management, customary practices, strengthening self capability of environmental and resource management in indigenous communities, and relationship building in conservation.

Transboundary Protected Areas: Making Peace With Nature – Melanie C. Berger

“[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?

Transboundary Protected Areas Worldwide (2007)

There were an estimated 227 Transboundary Protected Areas worldwide in 2007. (Schoon, n.d.)

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.


A sign indicating the partnership between The Ministry of Forest Economy (Government of Congo), the Congolaise Industrielle des Bois (CIB), and the Wildlife Conservation Society (photo: Wildlife Conservation Society).

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

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: 




World Parks Congress














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.

Clark, C. J., Poulsen, J. R., Malonga, R., Elkan Jr., P. W. (2009). Logging Concessions can extend the estate for Central African tropical forests. Conservation Biology, 23(5), 1281-93.

Kormos, C. F., Mittermeier, C. G., Gil, P. R., Sandwith, T., & Besancon, C. (2005). Transboundary conservation: a new vision for protected areas. Cemex.

McCallum, J., Schoon, M. (n.d.). Ecological benefits and costs of Transboundary Conservation Areas (TBCA). TBPA.com. Retrieved April 1, 2014, from http://www.tbpa.net/page.php?ndx=52

Sandwith, T., & Besancon, C. (2005). Trade-offs among multiple goals for transboundary conservation. Unpublished. http://theislamistsarecoming.wilsoncenter.org/sites/default/files/Besancon_Sandwith.pdf

Schoon, M. (n.d.). Brief History of Transboundary Protected Areas. TBPA.com. Retrieved April 1, 2014, from http://www.tbpa.net/page.php?ndx=17

Stokes, E. J., Strindberg, S., Bakabana, P. C., Elkan, P. W., Iyenguet, F. C., et al. (2010). Monitoring Great Ape and Elephant Abundance at Large Spatial Scales: Measuring Effectiveness of a Conservation Landscape. PLoS ONE, 5(4).

UNEP-WCMC. (2010). A-Z Guide of Areas of Biodiversity ImportanceUNEP-WCMC. Cambridge, UK. http://www.biodiversityA-Z.org.

United Nations Environment Programme

Wolmer, W. (2003). Transboundary Conservation: The Politics of Ecological Integrity in the Great Limpopo Transfrontier Park*Journal of Southern African Studies, 29:1, 261-278.

Wolmer, W. (2004). Tensions and paradoxes in the management of Transboundary Protected Areas. Policy Matters, 13, pp 137-146.

The role of people in conservation science – Charlie Hopkins

The role of “people” in conservation science
Charlie Hopkins

People are the leading cause and the solution for conservation. People are also a threat to conservation, the driving force behind conservation, victims of conservation efforts but also benefactors of conservation (Robertson & Hull, 2001). Although this initially seems paradoxical, the concept of “people” is vital in conservation science. Land-owners, indigenous people, developers, public and government all have diverse and significant roles to perform in every conservation project (Chan et al., 2007). Conservation is an active process, where actions must be to taken to change the dynamics of a given population; passive efforts seldom work in ecological restoration projects. Successful conservation strategies often involve a change to current human activity or removal of the population from the immediate area, because of the dichotomy between human industrialisation and ‘pristine wilderness’. Dualism might explain why conservation is in high current demand; positive public perception of these pristine reservations that are outside of their own daily environment drives increased conservation efforts. The tragedy of shifting baselines is a slippery slope in which future generations may not recognise significant declines in biodiversity due to the absence of exposure to such experiences. Conservation is a priority of modern people, preservation of nature being more of an issue now than previously, and it would be in our best interest to preserve our current ecosystems. The worst future venture for conservation science is the loss of public opinion; the conflict between economic development and environmental conservation needs to stop.

Environmental resources, such as interactive biodiversity or urban green spaces, are accessible through ‘top-down’ large organisations or governing bodies that have the financial and temporal investments available. Although top-down, reservation style conservation is achievable, and is possibly easier for the governing body without the community involvement, there are numerous socio-political advantages of having a biophilic population (Ban et al., 2013) including footprint awareness and respect when dealing with environmental implications of economic development.

Modern industrial populations that imprint a human foot print on the environment produce a demand for conservation. Essential primary industries such as agriculture are employed to feed populations, however often leave heavy footprints on water environments through practices such as water extraction, NPK fertiliser use, and pesticide use. Carbon based fuel extraction and consumption, which is primarily used for transport and electricity production, has implications on atmospheric integrity. Deforestation for timber, and accessibility to farmland, draining of wetlands and confinement of rivers have been the leading cause of terrestrial degradation while over harvesting of fish and other wildlife are the leading causes of ecological decline in marine settings. These activities produce vast economies across the globe, and conserving the species that these markets influence, is not as simple as reducing intensity. Virtually limitless complexity is involved in the social, cultural, economic and ecological characteristics of current conservation actions.

People also plan, initiate and execute conservation projects, although these projects are often biased toward human benefit through direct ecosystem services, cultural significance or logistical feasibility. Projects such as Genesis Energy Whio Recovery Program or Dulux New Zealand Kea Conservation Trust favour popular, well recognised species. Conservation efforts may become inclined toward species with a direct effect to a population or ecosystem service, when indirect services, such as soil organisms involved in decomposition, can be equally and vitally important (Berkes, 2004).

Conservation and biosecurity are dominant processes in New Zealand’s environmental management strategy. The extended period of isolation since New Zealand’s splitting from Gondwana has resulted in an advanced degree of biological differentiation and uniqueness, many of the native species are endemic and ‘living-fossils’. Species such as the Moa (Ratite), weta (Orthoptera) and Tuatara (Sphenodon spp.), along with many distinctive birds such as the Kakapo (Strigops habroptilus), Kaka (Nestor meridionalis) and Rifleman (Acanthisitta chloris) are exclusive to New Zealand biota. The evolution of such species in the absence of mammalian predators has led to the recent destruction of such fragile ecosystems over the last 700 years with the introduction of Ship rat (Rattus rattus), Norway rat (Rattus norvegicus), Kiore (Rattus exulans) and the house mouse (Mus musculus), Mustelidae, cats (Felis catus), possums (Trichosurus vulpecula),  among other pests.

New Zealand, like its biota, has a vast array of physical environments, densely compacted onto a small mainland causing steep environmental clines and niche partitions. The many partitions of native species distributions on off-shore islands has resulted in New Zealand’s conservation forces, primarily Department of Conservation and regional councils, to become world leaders in clearing islands of pest and weed species to create near pre-human environments for conservation efforts. The methods developed for pest control and native species conservation is starting to spill-over into mainland island projects such as Wainuiomata Mainland Island (GWRC) or Boundary Stream Mainland Island (DOC). Based on the current state of environment, the two largest aspects  of conservation practice in New Zealand are to improve pest and weed control methods while  arresting the decline of native species and also improving and maintaining public perception and education in regard to the environment (Logan, 2001).

New Zealand is rather lucky that ecological reconciliation projects can involve community groups, schools, or local companies “mucking-in”. In New Zealand, people are not so much harmed, but just restricted, by limitations imposed on economic quotas.  This is important in the New Zealand economy as the people are very biophilic, and the two largest industries, tourism and agriculture, heavily rely on environmental quality. These primary industries entrench a pride in the environment which is reflected in many local restoration projects such as Tiritiri Matangi Island in the Hauraki Gulf. Here, approximately 300,000 trees have been planted since the island became mammal pest free in 1993, and has since seen the reintroduction of North Island Robin, Takahe and Tuatara through the hard work of Department of Conservation & Supporters of Tiritiri Matangi Incorporated, including school and volunteer groups (Michel, Dickinson, Barratt, & Jamieson, 2010).  Despite these types of conservation ventures, conservation is often perceived to limit economic ventures such as logging, fishing, mining, and agriculture in New Zealand. Arresting New Zealand’s biodiversity decline would require economic change however this would, although perhaps not financially, return a gain on investment (McShane et al., 2011). Investment into the construction of mainland island projects such as Zealandia, which allow spill over or community visitation, would produce similar social and cultural benefits to near-shore island restorations such as Tiritiri Matangi or Kapiti Island.

Conservation in a global context is rather different to that in New Zealand, given that conservation is viewed by some as a luxury item. “Win-win” or “trade-off” approaches are described in conservation efforts in developing nations where wildlife is relied on for food or trade (McShane et al., 2011).  Ironically, the distribution of human poverty overlays areas of biological wealth (McShane et al., 2011) so  conservation efforts would  displace traditional hunting and conventional utilisation of biological material.  However, conservation can be done delicately with the establishment of non-traditional, reconciliation projects to compensate indigenous people, such as wildlife tour guides or scientific research assistants. Although this is sound in theory, the real world application of such a model may be limited as many funds from wildlife managers are dispersed through organisations before trickling down to the poorest groups that rely on, and have sacrificed, the resource. These indigenous groups are less removed from nature, and therefore are often victims of economic schemes of conservation projects.

Possibly the most important demand for conservation of biodiversity is the role of ecosystem services. We require more change because of the potential benefits of biodiversity, and human interaction with a healthy environment (Keniger, Gaston, Irvine, & Fuller, 2013). Ecosystem services such as climate regulation, crop pollination, toxin mediation, water and air purification are essential to a functional population (Balmford et al., 2002) which is achievable through reducing dualism and increasing environmental education and awareness. An example of biodiversity contributing to human benefit is the dissipation of tsunami waves on December 26, 2004 on beaches in the Indian Ocean that contained mangrove plantations were less damaged and recovered quicker than areas without (Danielsen et al., 2005).

Conservation for biodiversity can focus on two separate aspects: species diversity and functional diversity (Daily et al., 2000). Species diversity focuses on the variety of taxa found within an environment while functional biodiversity classes the utility of the community, such as colonising plants or top spatial predators. Biodiversity provides insurance for ecosystem services, through maintaining the benefit of knowledge and retaining genetic variation. Biodiversity also offers adaptability to environmental change and long term climate change. The value of ecosystem services is the amount that it saves the economy (Daily et al., 2000) however the value of biodiversity is much greater as  it includes social, cultural and ecological values. A diverse community will also lead to a broader, more intense demand for resources which makes it harder for exotic species to invade.

Modern development and industrialisation has led to a shift from a reliance on nature for subsistence hunting and gathering to a social and recreational approach (Keniger et al., 2013), although this has also led to a removal from nature and the development of a dualism between modern human populations and nature. Recently, post-war, it has become more recognised that direct interactions with nature have prominent advantages on human wellbeing and health. The primary influence of nature on human populations is no longer a resource for survival, rather an improved quality of life. Regular interactions with nature, such as recreating in green spaces, gardening or community-led conservation projects, can lead to increased cognitive function, physiological well-being, social participation and attitude and increased motivation or spiritual awareness (Keniger et al., 2013). Of course these are not only the direct effects on an individual; spill-over effects on a population could include decreased crime rates and anti-social behaviour, health benefits, growth in education rates and also increased economic productivity and an increased concern and awareness of nature.

It is in our best interest to conserve and interact with nature, rather than perceive the environment as an exploitable resource. Whether the intentions of conservation are focused on the benefits toward modern human populations or genuinely for preservation of biodiversity, community involvement in government supported projects will lead to healthier community and construct a sustainable future as a foundation for a better world.

A society grows great when old men plant trees whose shade they know they will never sit in.” – Greek proverb


Balmford, A., Bruner, A., Cooper, P., Costanza, R., Farber, S., Green, R. E., . . . Turner, R. K. (2002). Ecology – Economic reasons for conserving wild nature. Science, 297(5583), 950-953.

Ban, N. C., Mills, M., Tam, J., Hicks, C. C., Klain, S., Stoeckl, N., . . . Chan, K. M. A. (2013). A social-ecological approach to conservation planning: embedding social considerations. Frontiers in Ecology and the Environment, 11(4), 194-202.

Berkes, F. (2004). Rethinking community-based conservation. Conservation Biology, 18(3), 621-630.

Chan, K. M. A., Pringle, R. M., Ranganathan, J., Boggs, C. L., Chan, Y. L., Ehrlich, P. R., . . . Macmynowski, D. P. (2007). When agendas collide: Human welfare and biological conservation. Conservation Biology, 21(1), 59-68.

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Danielsen, F., Sorensen, M. K., Olwig, M. F., Selvam, V., Parish, F., Burgess, N. D., . . . Suryadiputra, N. (2005). The Asian tsunami: A protective role for coastal vegetation. Science, 310(5748), 643-643.

Keniger, L. E., Gaston, K. J., Irvine, K. N., & Fuller, R. A. (2013). What are the Benefits of Interacting with Nature? International Journal of Environmental Research and Public Health, 10(3), 913-935.

Logan, H. (2001). Gondwana invaded: an address on distinctive features of managing indigenous biodiversity in protected areas in New Zealand. Journal of the Royal Society of New Zealand, 31(4), 813-818.

McShane, T. O., Hirsch, P. D., Tran Chi, T., Songorwa, A. N., Kinzig, A., Monteferri, B., . . . O’Connor, S. (2011). Hard choices: Making trade-offs between biodiversity conservation and human well-being. Biological Conservation, 144(3), 966-972.

Michel, P., Dickinson, K. J. M., Barratt, B. I. P., & Jamieson, I. G. (2010). Habitat selection in reintroduced bird populations: a case study of Stewart Island robins and South Island saddlebacks on Ulva Island. New Zealand Journal of Ecology, 34(2), 237-246.

Robertson, D. P., & Hull, R. B. (2001). Beyond biology: toward a more public ecology for conservation. Conservation Biology, 15(4), 970-979.