Sequential Depletion of Australia’s Fisheries Resources: Ecosystem Effects and Sustainability
Emily Shaw
Griffith University Queensland Australia
2008
ABSTRACT
Global fisheries are in a state of crisis, which is of great concern both for the future sustainability of fisheries and also for marine ecosystems. One way in which fisheries impact marine ecosystems is through trophic effects, which arise from the removal of target species. Globally, there is a trend of ‘fishing down the food web’, where high trophic level stocks are depleted with subsequent shifts to lower trophic levels and this is expected to be unsustainable and to affect marine ecosystem functioning. The trends in mean trophic level of Australian fisheries from 1950 to 2005 were examined. The trend of fishing down the food web was not observed, but rather there was an increase in trophic level through the 1970s, 1980s and 1990s. Also, there was high variability in mean trophic level due to sequential harvesting, which is the overharvesting of one stock with sequential exploitation of other stocks. Indeed sequential harvesting was found to occur on a number of scales and is unsustainable and may result in ecosystem scale consequences. Future management needs to account for species interactions through the implementation of ecosystem based fisheries management.
1 INTRODUCTION
Marine fisheries have historically been viewed as an inexhaustible resource and have, therefore, been poorly regulated (Pauly et al., 2003). This has resulted in fisheries being largely non-sustainable and has led to a ‘tragedy of the commons’, in which each fisher aims to maximise their own gains (Hardin, 1968, Pauly et al., 2002). Globally, wild capture marine fisheries are in a state of crisis with 75% of stocks overexploited or fully exploited (Pauly et al., 1998; FAO 2007). Indeed global fish catches have been slowly declining since the end of the 1980s following an increase in fishing effort from the 1950s to 1970s, which led to the collapse of a number of major fisheries, for example the Peruvian anchovetta (Pauly et al., 2002). Today fisheries are still being mismanaged and increasing demand for seafood coupled with subsidy driven overfishing is leading to further unsustainable use of fisheries resources (Hilborn et al., 2003; Gewin, 2004). Furthermore, the expansion of the aquaculture industry may not reduce pressure on global capture fisheries but rather increase the pressure through the use of wild fish as feed for aquacultured species (Naylor et al., 2000).
Fishing has been found not only to impact on the species being targeted but also to impact on entire marine ecosystems (Coleman and Williams, 2002). Ecosystem effects from fishing occur primarily through bycatch of nontargeted species, impacts of fishing gear on habitats and through trophic effects caused by the removal of target species (Coleman and Williams, 2002; Pauly et al. 2003). These ecosystem effects from fishing, along with pollution, global climate change and invasive marine pests are the dominant factors believed to be threatening marine ecosystems and the biodiversity that they support (NRC, 1995). Of these processes, overfishing has historically occurred prior to other major forms of disturbance and has a profound effect on ecosystem structure and function (Jackson et al., 2001). Paleoecological, archaeological and historical data show that humans have overfished species to the point of ecological extinction, such that the species can no longer interact significantly within the community, causing marked changes in ecosystem structure (Jackson et al., 2001). Furthermore, these ecosystem structural changes may have time lags of decades to centuries (Jackson et al., 2001).
Overexploitation of species to the point of ecological extinction continues to occur with the sequential depletion of fish stocks (Hilborn et al., 2003). This occurs where a particular fishery is exploited to a point of economic extinction followed by the fishers moving to newly discovered stocks or to stocks that have become more accessible through technology (Hilborn et al., 2003). Pauly et al. (1998) found a distinct pattern to the sequential depletion of species whereby higher trophic level species are removed first, followed by a shift to lower trophic level species. This trend, known as fishing down the food web, has seen catches shift from large piscivorous fish to planktivorous fish and invertebrates (Pauly et al., 1998).
The pattern of removal of high trophic level species, or predators from a food web, might be expected to increase the abundance of lower trophic level prey species through a release from predatory control (Staneck, 1998). It may then be expected that with the decreasing mean trophic levels of global fisheries that there would be an increase in capture production. This, however, was not found to be the case, and may indicate that current exploitation regimes alter the structure of communities and are unsustainable (Pauly et al., 1998). Furthermore, the removal of predators from a food web may not simply lead to an increase in prey abundance, but rather lead to outbreaks of previously suppressed species that may be considered as pests (Pauly et al., 2002). For example, this is believed to have contributed to recent jellyfish blooms (Lynam et al., 2006).
Trophic cascades, resulting from interactions of high trophic level species with organisms occupying lower trophic levels within a food web, have been found to be widespread and numerous in the marine environment (Pace et al., 1999). Indeed it is the high trophic level species that may exert significant top down control over a community to an extent that they may be considered as keystone species (Paceet al., 1999). For example, a decline in sea otters, which are recognised as a keystone species in the North Pacific, initially from hunting and more recently from increased predation by killer whales, is believed to have resulted in an increased urchin population which then led to deforestation of kelp forests (Estes et al., 1998). Such trophic cascades, in conjunction with fishing down marine food webs, are believed to have reduced both the structural and functional diversity of marine ecosystems through a reduction in the number and length of pathways within food webs (Coleman and Williams, 2002; Pauly et al., 2002). This has then led to an increase in the variability of populations and a decrease in the resilience of ecosystems (Pauly et al., 2002).
Since the start of modern fisheries science in the 1950’s, governments have predominantly used single-species assessment models as the basis of fisheries management (Pauly et al., 2002). Presently there is a continued reliance on such models which do not take into account the effects that removing large numbers of individuals from a species will have on an ecosystem (Pauly et al., 2002). As such there has been a push from fisheries scientists to adopt ecosystem-based management tools to try to improve management of fisheries and ecosystems (eg. Pauly et al., 2002; Worm et al., 2002; Hall and Mainprize, 2004). For example, the use of marine protected areas (MPAs) as part of ecosystem-based management has been found to be capable of re-establishing trophic cascades where they had formerly collapsed due to fishing pressures (Shears and Babcock, 2002).
The need to restore and improve the sustainability of global fisheries was highlighted in 2002 at the World Summit on Sustainable Development in Johannesburg. All nations who attended the summit, including Australia, agreed to stop destructive fishing practices and establish marine protected areas and networks by 2012. Australia has developed an extensive network of MPAs relative to other parts of the world and may therefore be expected to be better able to preserve food web structure and ecosystem functioning. However, a number and range of destructive fishing practices still occur within Australia (Nevill, 2007). Destructive fishing practices may occur through adverse impacts on target species, such as overfishing, or by adversely affecting other species or ecosystems, for example through high levels of bycatch. Sequential depletion of fish stocks can be destructive in both of these ways, where the sequential overharvesting of target stocks can also be destructive to other species and ecosystems with which the target species interacts.
Despite the large area of Australia’s exclusive economic zone, Australian fisheries are relatively unproductive and contribute only a small fraction to the global capture production and therefore have little influence on global trends. The aims of this study are to examine the trends in mean trophic level of Australian fisheries from 1950 to 2005 and to compare these with the global trends. The ecosystem consequences resulting from a reduction of target species populations and the sustainability of Australia’s fisheries will also be reviewed, along with the implications for future management.
2 METHODS FOR TROPHIC LEVEL ANALYSIS
The methods used in this analysis were based on the methods of Pauly et al. (1998).
The mean trophic level for wild capture marine fisheries in Australia was determined for each year from 1950 to 2005 and trends in mean trophic level over time were examined. To do this, data on the total capture weight of each species for each year and the trophic level of each species was utilised. The catch statistics for Australia were obtained using the FAO’s FishStat Plus software (FAO Fisheries Department, Fishery Information, Data and Statistics Unit, 2000). This database has fisheries landing statistics based on reports from FAO member countries, including Australia, that are submitted annually. Cetaceans were excluded from this database as there is currently no commercial fishery for these species and the landings are measured by the number of individuals captured, and therefore capture biomass was not known. Catches of less than 0.5 tonnes are reported to the FAO as “less than 0.5 tonnes”, and were excluded from this analysis as the exact biomass caught could not be determined and catches of less than 0.5 tonnes were found to be insignificant in the analyses.
The trophic levels of each species were obtained from FishBase (Froese and Pauly, 2007) for fish species, and from the Sea Around Us database (Sea Around Us, 2007) for invertebrates. Trophic levels ranged from 2 to 4.6, where herbivorous species have a trophic level of 2 and the trophic level increases by one for each subsequent higher order of consumer (Pauly et al., 1998). Fractional trophic levels arose as a result of mixed diet composition of consumers (Pauly et al., 2000a). The trophic level data obtained from these databases was determined using both diet composition information and stable isotope analysis, along with modelled predictions using Ecopath software (Christensen and Pauly, 1992, Pauly et al., 2000b).
The mean trophic level was determined by the following equation:
Where mi is the mass of species i landed in a given year; Ti is the tropic level of species i; mT is the total mass of captures for a given year; and N is the number of species.
The initial analysis was performed with all Australian fisheries grouped together. However, subsequent analyses were also performed on individual FAO geographic regions. This was done because with advances in technology over time, fisheries have been able to move into new areas (for example, areas further offshore). This may have led to trends in mean trophic level occurring for each region and this effect may be masked if all areas are grouped together for analysis. The geographic regions that were analysed were the Indian Antarctic, Indian east, Pacific southwest and Pacific west central. Note that Australia also had catches in the Atlantic southwest; however, this was only in 1998 and 1999 and therefore no meaningful trend in trophic level over time could be established for this region.
3 RESULTS OF TROPHIC LEVEL ANALYSIS
The mean trophic level of Australian fisheries showed a decreasing trend in the 1950s and 1960s and an increasing trend from the 1970s until 2002, when trophic level began to decrease again (Figure 1). However, mean trophic level was highly variable with time (Figure 1). Much of this variation was found to be due to variability in catches of orange roughy, southern bluefin tuna and scallops. When these species were excluded from the analysis, the decreasing trend in mean trophic level in the 1950s and 1960s and the increasing trend in mean trophic level through the 1970s, 1980s and 1990s was still observed, however the variation in the data was reduced (Figure 2). These species had a large influence on the overall trends in mean trophic level due a combination of highly variable capture biomasses and their trophic levels, which were particularly large for orange roughy and southern bluefin tuna (trophic levels of 4.3 and 3.93, respectively) and particularly small for scallops (trophic level of 2).
Figure 1: Mean trophic level of Australian fisheries from 1950 to 2005
Figure 2: Mean trophic level of Australian fisheries from 1950 to 2005 with orange roughy, southern bluefin tuna and scallops excluded
When the trophic level analyses were performed for each geographical region it was again found that trophic levels were variable with time (Figure 3). Data on catches in the Indian Antarctic region are only from 1997 and are dominated by the Patagonian toothfish, which is a high trophic level species (Figure 3a). Mean trophic level in the Indian east region was relatively low as catches were dominated by invertebrates, particularly western rock lobster (Panulirus cygnus) and scallops (Figure 3b). There were several fluctuations in mean trophic level of this region, primarily due to variable catches of scallops, but also due to changes in catches of sharks, southern bluefin tuna, blue grenadier and sardines over time. The catches in the Pacific southwest region were dominated by fish and thus resulted in relatively high mean trophic levels, with orange roughy catches driving the large increase and subsequent decline in trophic level in the late 1980s to mid 1990s (Figure 3c). Catches in the Pacific west central region were dominated by low trophic level invertebrates, particularly penaeus shrimps and scallops (Figure 3d). The fluctuations in mean trophic level in this region were due to variability in catches of a number of species, for example scallops, penaeus shrimps and trochus shells.
Figure 3: Mean trophic level of Australian fisheries from 1950 to 2005 in (a) the Indian Antarctic region, (b) the Indian east region, (c) the Pacific southwest region and (d) the Pacific west central region
3.1 Comparison with Global Trends
The results of this trophic level analysis of Australian marine fisheries varies in a number of ways to the global trends. The results of Pauly et al.’s (1998) global analysis of mean trophic level show a trend of decreasing trophic level with time since the 1950s (Figure 4). At first, the global shift to lower trophic level species was found to allow for increased catches (Pauly et al., 1998). However, catches were then found to remain constant or decline over time, indicating that the current exploitation patterns in global fisheries were not sustainable (Pauly et al., 1998). However, the results for Australia show a general increasing trend in trophic level from the 1970s. Furthermore, the total capture production from marine fisheries in Australia has been increasing since 1950, with the exception of the 1990s where the catch remained fairly stable (Figure 5). Another difference is that the global trend in mean trophic level has little variation, with the exception of a period of lower trophic levels during the 1960s and early 1970s due to particularly large landings of low trophic level Peruvian anchovetta (Pauly et al., 1998). The mean trophic level in this analysis, however, was found to vary greatly with time.
Figure 4: Global trends in mean trophic level of marine fisheries landings, 1950 to 1994
Source: Pauly et al. (1998)
Figure 5: Total capture production of Australian marine fisheries from 1950 to 2005
Source data: FAO Fisheries Department, Fishery Information, Data and Statistics Unit, 2000
The high variability in mean trophic level for Australia may indicate that there have been dramatic shifts in the dominant species caught over time as a result of sequential depletion. This may have implications for the sustainability of Australia’s fisheries, despite the fact that fishing down the food web is not occurring and that total catch is not declining.
4 SEQUENTIAL DEPLETION IN AUSTRALIAN FISHERIES
4.1 Variation in Mean Trophic Level: An Indication of Sequential Depletion
Much of the temporal variation in mean trophic level was due to fluctuations in the catches of orange roughy, scallops and southern bluefin tuna (Figure 6). The catches of these species have fluctuated in the past due to overharvesting (see Box 1 for further information on the history and management of these fisheries). Despite the fact that catches of orange roughy, southern bluefin tuna and scallops have at times been very high and then had subsequent crashes, the total catch of Australian fisheries steadily increased until the 1990s (Figure 5). For this to have occurred there must have been subsequent increases in catches of other species following the crashes in the catches of orange roughy, southern bluefin tuna and scallops. Indeed further examination of catch trends for Australian fisheries reveal a trend of sequential depletion, where fisheries stocks are overexploited and subsequently fisheries shift to new species or locations. This trend of serial depletion can be seen on a number of scales.
Figure 6: Capture production of scallops, southern bluefin tuna and orange roughy in Australian fisheries
Source data: FAO Fisheries Department, Fishery Information, Data and Statistics Unit, 2000
Box 1: History and Management of 3 Australian Fisheries
Orange Roughy Fishery
The orange roughy fishery in Australia has been established relatively recently, with commercial fishing beginning in 1986 (Kailola et al., 1993). The first records of orange roughy in Australian waters are from 1972 but commercial fishing did not commence until the discovery of large aggregations of orange roughy in 1986 (Kaiolola et al., 1993). The fishery is largely based on targeting aggregations, where catches rose to a maximum of approximately 40, 000 tonnes by 1990 (Kailola et al., 1993). These exploitation rates were unsustainable and catches rapidly declined (Figure 6). The orange roughy fishery is Commonwealth managed in the Southern and Eastern Scalefish and Shark Fishery (SESSF), South Tasman Rise Fishery and in the Cascade Plateau, where their current status is overfished in both the SESSF and South Tasman Rise Fishery (Larcombe and McLoughlin, 2007).