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- Materials and methods
Protected areas are a mainstay of biodiversity conservation throughout the world, but their performance has not been adequately assessed (Gaston et al. 2008). Uncertainty is particularly high in the case of freshwater biodiversity, with even the representation of freshwater features in protected areas being poorly known (Herbert et al. 2010). Doubts about the benefits of existing protected areas for freshwater conservation arise because few such areas have been designed or managed with freshwater biodiversity specifically in mind (Saunders, Meeuwig & Vincent 2002; Abell, Allan & Lehner 2007). For example, terrestrial protected areas often include only part of a river's catchment or use a river as a boundary rather than fully including it (Nel et al. 2007). Consequently, rivers within protected areas are often vulnerable to transmission of impacts from land and water use beyond reserve boundaries (Pringle 2001). In addition, fresh waters within protected areas may be stocked with or invaded by alien fish (Pittock, Hansen & Abell 2008).
Declines and losses of freshwater biodiversity within protected areas have been widely documented. For example, Barendregt, Wassen & Schot (1995) attributed a loss of wetland vegetation diversity in a Dutch reserve to extrinsic influences such as altered hydrology and nutrient enrichment. Disappearances of frogs ascribed to chytrid fungus infections and species losses associated with fish introductions have also been reported within protected areas (Schindler 2000; Skerratt et al. 2007).
Do protected areas nonetheless perform better than comparable unprotected areas in maintaining freshwater species? Only a few, mostly small-scale, studies have tackled this question, and they have produced mixed results. Positive findings include those of Baird & Flaherty (2005), who observed that villagers in Lao PDR reported increased fish abundance after protection zones were established, and Cucherousset et al. (2007), who found that eels were more abundant and larger in protected portions of a French wetland than in fished areas. Similarly, Sanyanga, Machena & Kautsky (1995) reported that the mean body size of commercial species was larger in protected than in fished areas of Lake Kariba, Zimbabwe. On the other hand, Mancini et al. (2005) observed that a macroinvertebrate index did not differ significantly between stream reaches inside and outside of Italian protected areas, and Srinoparatwatana & Hyndes (2011) found no consistent differences in abundance or biomass of wetland fish species between a protected area and an adjacent fished area in Thailand.
Comparisons of biodiversity within and beyond protected areas are a logical way to assess performance, but may be confounded if the places being compared differ in natural environmental features (Mas 2005). Such confounding is unfortunately likely because the creation of reserves often favours land with little potential for agricultural, urban or industrial use, such as steeply sloping terrain with low soil fertility (Scott et al. 2001; Pressey et al. 2002). Confounding by extraneous variables that covary with those of concern is a widespread problem when observational studies of the natural environment attempt to assign causes to observed differences or changes (Beyers 1998). In such studies, the role of potentially confounding factors cannot be balanced or minimized as it would in controlled experiments, where treatments are assigned randomly to experimental units and extraneous factors are held constant while those of interest are varied. However, propensity scores (Rosenbaum & Rubin 1983) provide a way to adjust comparisons for the influence of confounding factors in observational studies. A propensity score is the probability that an observational unit has received a specified treatment given the unit's values of potentially confounding factors. A comparison that is not biased by the confounding factors can be achieved by choosing observational units such that propensity scores are balanced among treatments.
I used propensity scores to evaluate how protected area status affects riverine fish assemblages over a large spatial extent: Australia's Murray–Darling Basin, which covers 1 061 469 km2 or about one-seventh of the continent. Fish are an appropriate focus for a study of protected area performance because they are prominent in the precipitous decline of freshwater biodiversity that has been reported around the globe (Dudgeon et al. 2006; Strayer & Dudgeon 2010). Nearly half of the world's known fish species live in fresh waters (Lévêque et al. 2008), and 37% of those evaluated by the IUCN in 2008 were assessed as threatened (Vié, Hilton-Taylor & Stuart 2009). The Murray–Darling river system is also a salient case study, because many of its indigenous fish species have suffered severe historical declines (Cadwallader 1978) and several are formally listed as threatened (Koehn & Lintermans 2012). Current efforts to sustain the river system's biodiversity while accommodating social and economic needs for water illustrate challenges that other parts of the world are also likely to face (Pittock & Connell 2010).
Using data from a basin-wide monitoring programme, I first compared fish assemblages between all sampling sites located within protected areas (hereafter ‘protected sites’) and all sites outside of such areas (‘unprotected sites’). I then used propensity scores to select a subset of protected and unprotected sites that were matched for natural environmental features and repeated the comparison. I hypothesized that once the effect of confounding factors was controlled, native fish would be more diverse and abundant in protected areas because such areas would provide at least some shelter from threatening processes associated with anthropogenic disturbance of rivers and their watersheds.
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- Materials and methods
The sampling sites were most concentrated in the south and east of the basin where stream density is highest, and 69 (8·2%) were assessed as being within protected areas (Fig. 1). Protected sites were distributed across all IUCN protected area management categories apart from category VI (Table 1), but most fell in category II, that is, large natural or near-natural tracts reserved to protect characteristic species and ecosystems and broad-scale ecological processes, while allowing for compatible human uses (Dudley 2008). Almost half were within extensive, mountainous national parks: Kosciuszko National Park in New South Wales (6900 km2) and the adjoining Alpine National Park in Victoria (6600 km2) and Namadgi National Park in the Australian Capital Territory (1060 km2). Nearly a quarter were in strip reserves along riverine corridors. On average, protected sites lay farther south, had smaller catchments and steeper slopes and were colder than unprotected sites (t-tests; P ≤ 0·001). However, the two site types did not differ significantly in mean longitude (P = 0·160) or distance from the mouth (P = 0·113; Fig. 2).
Table 1. Percentage of protected sites within each IUCN protected area management category
|IUCN category||% of sites|
Figure 2. Mean values of environmental variables (±95% confidence limits) for all protected (P) and unprotected (U) sites and for an environmentally matched subset of sites. Values for catchment area and slope were calculated from logarithmically transformed data.
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The fish data included 36 described species, two undescribed species of Galaxias and the genus Hypseleotris, comprising H. klunzingeri plus undescribed species (Table 2). On average, protected sites had significantly lower native fish richness (Poisson test; P < 0·001) and abundance (t-test; P = 0·015) than unprotected sites, but the two site types did not differ significantly in richness or abundance of alien fish (P > 0·05; Fig. 3). Six fish taxa were significantly more abundant at protected sites, and another six were significantly more abundant at unprotected sites (Mann–Whitney U-tests; Table 2).
Table 2. Mean number of individuals of each fish taxon collected at all protected and unprotected sites and at an environmentally matched subset of sites
|Taxon||All sites||Matched sites|
|Ambassis agassizii (T)||0·043||0·006||0·046||0·005|
|Bidyanus bidyanus (T)||0·022||0·073||0·023||0·015|
|Carassius auratus (A)||1·109||2·134||1·177||1·070|
|Carassius carassius (A)||<0·001||0·002||<0·001||<0·001|
|Craterocephalus amniculus (T)||<0·001||0·103||<0·001||0·160|
|Craterocephalus fluviatilis (T)||<0·001||0·001||<0·001||0·010|
|Craterocephalus stercusmuscarum fulvus (T)||1·000||3·471||1·062||2·955|
|Cyprinus carpio (A)||12·261***||9·173***||13·015||4·535|
|Gadopsis bispinosus (T)||5·152***||1·458***||5·469**||1·980**|
|Gadopsis marmoratus (T)||0·261*||1·129*||0·246*||0·885*|
| Galaxias brevipinnis ||0·058||0·017||0·062||0·090|
|Galaxias fuscus (T)||<0·001||0·103||<0·001||0·790|
| Galaxias maculatus ||0·681*||0·078*||0·246||0·310|
|Galaxias olidus (T)||3·341||6·019||2·992||8·865|
|Galaxias sp. 1||<0·001*||1·793*||<0·001**||2·295**|
|Galaxias sp. 2||0·188||0·231||0·200||0·115|
|Galaxias truttaceus (T)||<0·001||0·003||<0·001||<0·001|
|Gambusia holbrooki (A)||3·145***||35·348***||3·338||18·225|
| Leiopotherapon unicolor ||10·862||1·370||11·531||0·845|
|Maccullochella macquariensis (T)||0·116||0·052||0·123||0·090|
|Maccullochella peelii (T)||0·630||0·938||0·669||0·485|
|Macquaria ambigua (T)||1·442||1·386||1·531||0·565|
|Macquaria australasica (T)||0·014||0·012||0·015||0·010|
|Melanotaenia fluviatilis (T)||0·884||1·821||0·938||1·355|
|Misgurnus anguillicaudatus (A)||0·014||0·073||0·015||<0·001|
|Mogurnda adspersa (T)||<0·001||0·008||<0·001||<0·001|
|Nannoperca australis (T)||0·043||1·856||0·046||4·130|
| Nematalosa erebi ||11·543**||26·612**||12·254||8·130|
| Neosilurus hyrtlii ||0·159***||0·027***||0·169*||0·060*|
|Onchorhynchus mykiss (A)||8·174***||0·949***||8·231***||3·450***|
|Perca fluviatilis (A)||0·957||4·531||1·015||1·750|
| Philypnodon grandiceps ||2·703||1·473||2·869||2·285|
| Philypnodon macrostomus ||0·014||0·009||0·015||<0·001|
| Retropinna semoni ||2·399*||5·772*||2·546||3·210|
|Rutilus rutilus (A)||<0·001||0·038||<0·001||0·010|
|Salmo trutta (A)||7·732***||1·916***||6·885*||4·070*|
|Tandanus tandanus (T)||0·080||0·213||0·085||0·010|
|Tinca tinca (A)||<0·001||0·147||<0·001||0·785|
Figure 3. Mean values of fish assemblage attributes (±95% confidence limits) for all protected (P) and unprotected (U) sites and for an environmentally matched subset of sites. Values for abundance of native and alien species were calculated from logarithmically transformed data.
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The logistic regression model was highly significant (P ≪ 0·001) and assigned mean propensity scores of 0·18 for sites that actually were within reserves and 0·07 for sites that were not. The matching process selected 65 of the protected sites and 100 of the unprotected ones. The environmental differences between protected and unprotected sites were much smaller for this matched subset than for all of the sites (Fig. 2), and none of them was statistically significant (t-tests; P > 0·05). Native richness, native abundance and alien abundance did not differ significantly for the matched subset (Poisson and t-tests; P > 0·05), but alien richness was significantly higher in protected sites (P = 0·033; Fig. 3). Significant differences in abundance remained after matching for six fish taxa, four of which were more abundant within protected areas (two of them native), while two were more abundant outside (both native; Mann–Whitney U-tests; Table 2).
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On average, fish sampling sites within protected areas were on smaller, colder and steeper streams than unprotected sites. A bias in allocation of protected status towards higher elevations has been observed in many parts of the world (Keith 2000; Scott et al. 2001; Januchowski-Hartley et al. 2011) and may reflect the scenic value of mountainous terrain and its lower suitability for competing land uses such as agriculture. The Murray–Darling Basin is straddled by extensive alpine and subalpine reserves but also has some large protected areas at low elevations. However, the latter lie mainly in arid and semi-arid tracts where non-ephemeral streams are scarce. Consequently, few of the lowland sites but many of the upland ones fell within protected areas.
The existence of statistically significant natural environmental differences between sites inside and outside of reserves precluded inferring benefits of protected status from a universal comparison of protected and unprotected sites. However, the use of propensity scores enabled a comparison that was geographically and environmentally balanced and suggested that protected status had little overall effect on fish assemblages. Two native species (plus two alien species) were significantly more abundant in protected sites than in matched unprotected sites, but two native species were significantly less abundant, and given that 39 taxa were considered, two significant differences are expected by chance at α = 0·05.
Beneficial effects of the reserve system were perhaps under-estimated by this analysis because movements of fish out of protected areas could have spread benefits from protected to unprotected areas. However, any fish that moved randomly among the sampling sites in this study would have spent only about 8% of its time in reserves, and consequently, any benefit received from reserves by highly mobile species is likely to be limited. In addition, protected areas might have provided benefits to fish living in downstream, unprotected areas, for example by influencing water quality. However, if this effect were large, it should have been manifest in assemblage differences between protected and unprotected sites, because on average, the proportion of a site's catchment that was within protected areas was about twice as great for protected as for unprotected sites.
Protected area status probably has little effect on fish assemblages in the Murray–Darling Basin because it does not, by itself, exclude some major anthropogenic influences. For example, reductions in river flows and the inundation of floodplain wetlands, caused by impoundment and water abstraction, have been implicated as a principal cause of native fish decline in the Murray–Darling Basin (Cadwallader 1978). The flow (and temperature) regimes of rivers in the basin's protected areas can be altered by impoundment and abstraction in upstream, unprotected river reaches or even by within-reserve infrastructure such as hydro-electric developments (Kingsford 2000). Water chemistry within protected areas can be affected by upstream land use or within-reserve tourism developments (Pickering, Harrington & Worboys 2003). Protected status also does not prevent fish harvesting, because angling is generally permitted within reserves, or exclude alien fishes, which impose a great constraint on native fish conservation. Introduced salmonids, especially Salmo trutta and Oncorhynchus mykiss, have severely reduced the abundance and geographical ranges of native Galaxias species in south-eastern Australia (Crowl, Townsend & McIntosh 1992; McDowall 2006), limiting the galaxiids of some protected areas to fragmented populations with a high risk of genetic diversity loss through local extinction (Green 2008). Other alien fishes in the basin with well-documented adverse impacts include Cyprinus carpio (Koehn 2004) and Gambusia holbrooki (Pyke 2008). The results presented here confirm that alien fishes are plentiful in protected areas and even suggest that some alien species are more abundant within such areas than outside of them.
Studies reporting benefits of protected areas for freshwater fish species have typically involved specific provisions for fish conservation such as ‘no-take’ zones (Sanyanga, Machena & Kautsky 1995; Baird & Flaherty 2005; Cucherousset et al. 2007). In the case of the Murray–Darling Basin, environmental water allocations are already being used to enhance spawning and recruitment of native fish in some protected areas (Tonkin, King & Mahoney 2008; Rayner, Jenkins & Kingsford 2009; King et al. 2010) and might become a widespread and routine management tool in protected areas where river flows and wetland inundation are affected by water resource development or climate change. Opportunities also exist to reduce alien fish populations within the basin's protected areas. Broad-scale eradication of trout would not be socially acceptable because of their recreational popularity and economic value (Jackson et al. 2004), but local removal is technically feasible (Lintermans 2000) and can be done selectively to protect threatened native species (Raadik, Saddlier & Koehn 1996). Some other alien fishes such as C. carpio and G. holbrooki are more difficult than trout to control or eradicate, but have little social or economic value and therefore might be widely targeted within protected areas.
Many authors have advocated the creation of new types of protected areas dedicated specifically to freshwater conservation (Moyle & Yoshiyama 1994; Saunders, Meeuwig & Vincent 2002; Kingsford & Nevill 2005; Abell, Allan & Lehner 2007; Suski & Cooke 2007; Humphries & Winemiller 2009; Williams et al. 2011). In the Murray–Darling Basin, several intensively managed ‘demonstration reaches’ have been established as part of a Native Fish Strategy that aims to rehabilitate indigenous fish populations (Koehn & Lintermans 2012). These reaches do not have formal reserve status but might serve as a precursor to freshwater protected areas by fostering community awareness and support (Barrett & Ansell 2003). Other potential candidates for freshwater protected areas in the Murray–Darling Basin include habitats of threatened fish species that are poorly represented in existing reserves and drought refuges, which are likely to become increasingly important in the light of projected long-term climatic drying (Leblanc et al. 2012).
Much remains to be done to develop an adequate system of freshwater protected areas for Australia (Fitzsimons & Robertson 2005; Stein & Nevill 2011) and the world (Abell, Allan & Lehner 2007). The actions necessary to achieve this aim include the creation of new reserves to fill gaps in coverage of the full spectrum of freshwater species and ecosystems and to embrace critical habitats such as refuges. However, the present study provides broad-scale evidence to reinforce concerns that protected status per se will not adequately conserve freshwater biodiversity, especially where the principal threats are not habitat destruction or direct harvesting but more insidious and pervasive factors such as alien species and climate change (Abell, Allan & Lehner 2007; Pittock, Hansen & Abell 2008; Rahel, Bierwagen & Taniguchi 2008; Williams et al. 2011). Area protection will achieve its potential for freshwater conservation only if coupled with intensive management to abate threats. For the Murray–Darling Basin, technical advances in control of alien fishes (Britton, Gozlan & Copp 2010) and resolution of competing demands for water (Connell & Grafton 2011) are likely to be critical to enabling protected areas to maximize their contribution to freshwater conservation. The same is likely to hold true in other parts of the world where invasive fishes are rife or dry climates intensify pressures on water resources.