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
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).
2
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 (Pace et 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 singlespecies 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
3
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.
4
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:
mi  Ti
mT
iN
MeanTrophicLevel  
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
5
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.
6
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).
3.45
Mean Trophic Level
3.4
3.35
3.3
3.25
3.2
3.15
3.1
3.05
3
2.95
1950
1960
1970
1980
1990
2000
Year
Figure 1: Mean trophic level of Australian fisheries from 1950 to 2005
7
3.45
Mean Trophic Level
3.4
3.35
3.3
3.25
3.2
3.15
3.1
3.05
3
2.95
1950
1960
1970
1980
1990
2000
Year
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.
8
(b)4.2
4
4
3.8
3.8
3.6
3.6
Trophic Level
Trophic Level
(a) 4.2
3.4
3.2
3.4
3.2
3
3
2.8
2.8
2.6
1950
1960
1970
1980
1990
2000
2.6
1950
1960
1970
Year
4.2
(d) 4.2
4
4
3.8
3.8
Trophic Level
Trophic Level
(c)
3.6
3.4
3.2
1990
2000
1990
2000
3.6
3.4
3.2
3
3
2.8
2.8
2.6
2.6
1950
1980
Year
1960
1970
1980
1990
2000
1950
Year
1960
1970
1980
Year
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.
9
Figure 4: Global trends in mean trophic level of marine fisheries landings, 1950 to 1994
Source: Pauly et al. (1998)
300,000
Catch (tonnes)
250,000
200,000
150,000
100,000
50,000
19
50
19
54
19
58
19
62
19
66
19
70
19
74
19
78
19
82
19
86
19
90
19
94
19
98
20
02
0
Year
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.
10
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.
45000
40000
35000
Catch (tonnes)
30000
25000
Scallops
Southern Bluefin Tuna
Orange Roughy
20000
15000
10000
5000
0
1948
1958
1968
1978
1988
1998
Year
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
11
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).
Southern Bluefin Tuna Fishery
Southern bluefin tuna is a migratory species in the Southern Hemisphere that has been fished commercially
in Australia since the 1950s (Kailola et al., 1993). The fishery began based on pole and line fishing in the
1950s, where the range and efficiency of the fishery expanded in the 1960s through the use of aerial spotting
(Kailola et al., 1993). Purse seining for southern bluefin tuna in Australia began in the 1970s (Kailola et al.,
1993). Southern bluefin tuna are also fished in Australian waters by Japanese longliners (Kailola et al.,
1993).
In Australia the southern bluefin tuna catch peaked in 1982 at approximately 22, 000 tonnes and has since
declined to approximately 5, 000 tonnes (Figure 6). This decline in catch has been due to a reduction in
spawning stock which resulted from excessive catches from the 1960s through to the 1980s (Findlay, 2007).
The southern bluefin tuna fishery is internationally managed through the Commission for the Conservation
of Southern Bluefin Tuna. Under this management system each member country, and some non-member
countries, are allocated proportion of a global total allowable catch (TAC). Australia’s TAC is approximately
5, 000 tonnes, however it is expected that the potential sustainable yield could be much higher than this if
spawning stocks were allowed to rebuild (Findlay, 2007).
Scallop Fishery
The major scallop fishery in Australia is the southern scallop fishery. This fishery is Commonwealth
managed and its current status is overfished (Larcombe and McLoughlin, 2007). In the early 1900s dredging
for scallops began in eastern Tasmania (McLoughlin, 2007). By the 1970s, when all of the eastern
Tasmanian grounds were depleted, the operations moved to Victoria where new scallop beds were found and
12
then to Bass Strait where new beds were found again (McLoughlin, 2007). This sequential depletion of
scallop beds has resulted in fluctuating ‘boom bust’ catches of scallops (Figure 6, McLoughlin, 2007). Other
catches of scallops in Australia are of saucer scallops in northern Australia, where it is expected that the
density of scallops in a number of beds in this fishery had been reduced from one animal per m 2 to one
animal per 150 m2 by the late 1980s (Kailola et al., 1993). Management of saucer scallop fisheries is
currently undertaken by state governments.
4.2 Sequential Depletion
4.2.1 Sequential Depletion at the National Scale
At a national scale the unsustainable harvest of species and subsequent shift to new species can be observed
in capture production trends. If sequential depletion was not occurring then the relative contribution of each
species or group to the total catch should remain fairly constant over time. However, this was not found to be
the case for Australian fisheries. Several species that have been major contributors to the total catch of
Australian fisheries have been sequentially overharvested throughout the development of Australian
fisheries, with examples of the dominant species given below.
Australia’s total capture production increased rapidly in the 1970s and 1980s (Figure 5) and associated with
this increase were marked increases in the landings of penaeus shrimps, scallops and southern bluefin tuna
(Figure 7a). The catches of these species peaked in the 1980s but were found to be unsustainable and
subsequently declined or were reduced through the implementation of quotas (Figure 7a; Findlay, 2007;
McLoughlin, 2007; Stobutzki and McLoughlin, 2007). Following the decline in catches of these species in
the late 1980s there were subsequent increases in catches of other species. Notably, the orange roughy
fishery was rapidly expanded, there were large catches of silver gemfish and yet another peak in scallop
catches (Figure 7b). Again the increase in catches of these species was found to be unsustainable to such an
extent that all three of these species still have a current status of overfished (Larcombe and McLoughlin,
2007). Later in the 1990s there was increased effort in targeting blue grenadier, yellowfin tuna and warehou
(Figure 7c). By early this century catches of these species had declined to an extent where yellowfin tuna and
blue grenadier are currently classified as overfished (Larcombe and McLoughlin, 2007). Capture biomass in
Australia is currently dominated by catches of clupeoids from the rapidly expanding South Australian
sardine fishery (Figure 7d). Time will tell as to whether this species can be harvested sustainably or whether
it will be yet another fishery that is sequentially overharvested.
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(a)
40000
35000
45000
(b)
Southern bluefin tuna
Scallops
Penaeus shrimps
40000
35000
30000
30000
25000
Catch (tonnes)
Catch (tonnes)
Silver gemfish
Scallops
Orange roughy
20000
15000
25000
20000
15000
10000
10000
5000
0
1940
5000
1950
1960
1970
1980
1990
2000
0
1940
2010
1950
1960
1970
Year
(c)
10000
9000
1980
1990
2000
2010
Year
70000
(d)
Blue grenadier
Yellowfin tuna
Warehou nei
Clupeoids
60000
8000
50000
6000
Catch (tonnes)
Catch (tonnes)
7000
5000
4000
3000
40000
30000
20000
2000
10000
1000
0
1940
1950
1960
1970
1980
Year
1990
2000
2010
0
1940
1950
1960
1970
1980
1990
2000
2010
Year
Figure 7: Sequential harvesting of species (a) in the 1980s, (b) in the early 1990s, (c) in the late 1990s, and
(d) in the 2000s
Source data: FAO Fisheries Department, Fishery Information, Data and Statistics Unit, 2000
4.2.2 Sequential Depletion at the Multispecies Fishery Scale
The sequential depletion of stocks can also been seen at the scale of individual multispecies fisheries. An
example of this is the southern and eastern scalefish and shark fishery (Commonwealth trawl sector). The
fishery began in the early 1900s and was dominated by stream trawlers prior to 1950, by Danish seiners from
the early 1950s to early 1970s and then by modern trawlers until present (Morison et al., 2007). When the
fishery began, trawlers targeted continental shelf species in less than 200 m depth (Morison et al., 2007).
Prior to 1930, tiger flathead was the dominant species caught, however, a stock decline in the late 1940s led
to increases in the catches of redfish and jackass morwong (Morison et al., 2007). In the 1950s Danish
seiners expanded the range of their operations from New South Wales south to Victoria (Morison et al.,
2007). Similarly, trawlers shifted to fish more southerly and deeper waters in the 1970s (Morison et al.,
2007). In the 1970s and 1980s there was a large increase in effort and capacity in the fishery which led to a
worsening economic state, which was exacerbated by crashes of stocks, such as the eastern gemfish
(Morison et al., 2007). In the late 1980s, the development of the orange roughy fishery led to a further
increase in capacity and effort (Morison et al., 2007). In this fishery alone, orange roughy, deepwater sharks,
blue warehou, eastern gemfish, redfish, smooth oreo dory and other oreo dories all have a current status of
14
overfished (Larcombe and McLoughlin, 2007). This case study demonstrates that without adequate
management controls on a fishery, both species and areas will be sequentially depleted.
4.2.3 Sequential Depletion at the Single Species Fishery Scale
The sequential depletion of stocks can also occur on a finer scale of a single species fishery. A prime
example of this is the southern scallop fishery. This fishery is for a single species, the southern scallop,
Pecten fumatus (McLoughlin, 2007). Individual beds of scallops have been sequentially depleted in this
fishery, where beds were dredged to a point where they were no longer economically viable and then new
beds were sequentially discovered and overfished (McLoughlin, 2007). This process was repeated until a
point where there were virtually no productive scallop beds remaining (McLoughlin, 2007).
4.3 Summary of Sequential Depletion and Trophic Level Trends
Pauly et al. (1998) recognised that current trends in global fisheries were unsustainable by identifying a trend
of decreasing mean trophic level, initially with an increase in capture biomass and then a subsequent decline
in capture biomass. In this analysis it has been found that mean trophic level has increased in Australia
through the 1970s, 1980s and 1990s, with increase in capture biomass from the 1950s to the end of the 1980s
and a constant capture biomass in the 1990s. In most cases, this has not been able to occur due to sustainable
management of Australia’s fisheries, but rather has occurred due to the unsustainable sequential
overharvesting of stocks.
Capture biomass was able to increase from the 1950s to the end of the 1980s due to increased effort and
efficiency through improved technology (Kailola et al., 1993). Fisheries were able to expand further offshore
and to greater depths during this time and this contributed to the increase in capture biomass (Kailola et al.,
1993). This may also explain why trophic level has increased, as most of the species being targeted offshore
and at depth, for example orange roughy and Patagonian toothfish, have high trophic levels. By the 1990s
there were few areas that had not been exploited and this may explain why capture biomass remained
constant and did not continue to increase in this decade. Capture biomass started to increase again from
2002. This was largely driven by rapid increases in catches in the South Australian sardine fishery. Because
this species is of a low trophic level, a resulting decrease in trophic level since 2002 has also been observed.
This scenario resembles the global trends at the point where a shift to lower trophic level species resulted in
an initial increase in capture biomass. Whether this increase in capture biomass will be sustainable or
whether Australian fisheries capture biomass will decline, as was seen in the global trends, will be seen in
the future.
15
5
ECOSYSTEM EFFECTS OF SEQUENTIAL
HARVESTING
The sequential harvesting of Australia’s fisheries is not just a concern with regard to their long-term
sustainability, but is also a concern as to the ecological consequences that arise from removal of a large
proportion of a population. Although fishing operations generally cease at a point of commercial extinction,
where it is no longer economically viable to continue to operate in that fishery, the total extinction of species
rarely occurs from fishing (Hilborn et al., 2003). It is more common that species should become ecologically
extinct, where their population will no longer be of sufficient size to interact significantly within the
community or ecosystem (Jackson et al., 2001). This causes simplification of food webs through a reduction
in number and length of food chains and results in decreased ecosystem resilience and increased variability
of individual stocks (Coleman and Williams, 2002, Pauly et al., 2002). Furthermore, the effects of removing
a large proportion of a population may not become apparent until decades or even centuries after the
population decline and this time lag can make it difficult to quantify or even detect the trophic effects of
sequential harvesting (Jackson et al., 2001).
The best studied trophic effects of removal of large numbers of a single species are generally of high trophic
level keystone species. Perhaps the best known example of this is of the ecosystem effects of removing sea
otters in the North Atlantic (Estes et al., 1998). Whilst not all high trophic level species are recognised as
keystone species, the removal of a large proportion of their population would still be expected to affect the
community with which they were interacting. For example, phytoplankton in the North Pacific are more
abundant when high trophic level pink salmon are also more abundant (Shiomoto et al., 1997). This is
believed to be from a reduction in predation on phytoplankton due to increased predation on
macrozooplankton by pink salmon when they are more abundant (Shiomoto et al., 1997). This raises
questions as to what ecological changes may be occurring in Australia due to overharvesting of high trophic
level species such as orange roughy or southern bluefin tuna.
It is not just the removal of high trophic level species that can impact marine ecosystems, but also lower
trophic level species. The effects of removing lower trophic level species have predominantly been studied
with respect to the consequences for their predators, where fluctuations and reductions in prey populations
adversely affect predator populations (Hall, 1999a). There have been numerous studies linking declines in
piscivorous seabird populations to reductions in populations of their prey that have resulted from fisheries
for these prey species (Tasker et al., 2000). For example, the breeding success of African penguins
(Spheniscus demersus) has been found to be dependent on the abundance of sardines (Sardinops sagax) and
anchovy (Engraulis encrasicolus) (Crawford et al., 2006). These species are prey items for the African
penguin and their abundance is markedly altered by fishing pressure (Crawford et al., 2006). This raises
concerns about potential ecosystem consequences of the rapidly expanding sardine fishery in South Australia
16
(Figure 7d). Little scientific research has been undertaken to assess the ecosystem effects of reducing the
biomass of sardines in Australia, however, they are believed to be a major food source for juvenile southern
bluefin tuna, as well as numerous species of seabirds and mammals (Shanks, 2005).
Depletion of any population, regardless of trophic level, could be expected to contribute to ecosystem-scale
consequences. This is particularly the case where the exploited species is considered to be an ecosystem
engineer. Ecosystem engineers enhance the biodiversity and functioning of an ecosystem either through their
morphological or behavioural traits that create a more complex habitat (Coleman and Williams, 2002).
Examples of such organisms can range from bivalve molluscs, which increase the physical complexity of
habitats and alter water column and sediment organic matter content; to sea cucumbers, which alter the
oxidation status of the sediment; to sea turtles and dugongs, which alter seagrass bed habitats (Coleman and
Williams, 2002).
In Australia several species which may have had roles as ecosystem engineers have been overexploited.
These include sea cucumbers prior to European settlement and dugongs, turtles and pearl oysters after
European settlement (Jackson et al., 2001). The dugongs, pearl oysters and turtles were particularly heavily
exploited where populations collapsed and have still not regained more than a fraction of their original
abundance (Jackson et al., 2001). Indeed Jackson et al. (2001) suggest that the loss of seagrass in Moreton
Bay can be attributed, at least in part, to the marked decline in sea turtles, where cropping by turtles reduces
the flux of nutrients and organic matter that reaches sediments and ultimately reduces the vulnerability of
seagrasses to disease. The exact ecosystem consequences of overharvesting species that were exploited
during early European settlement and prior to European settlement in Australia may be difficult to discern as
ecological records that date back that far are rare. In the future the effects of more recent cases of
overexploitation of marine species may become apparent.
17
6 FUTURE MANAGEMENT REQUIREMENTS
Overfishing both in Australia and globally has resulted in depletion of stocks, which could result in
ecosystem-scale consequences. To successfully manage fisheries and to minimise broader ecosystem-scale
effects it has been suggested that ecosystem based fisheries management be adopted (Hall and Mainprize,
2004, Pauly et al., 2002, Worm et al., 2002). Ecosystem based management should take into account species
interactions and the consequences for communities and ecosystems that arise from removing target species
(Hall, 1999b). It has also been suggested that, rather than aiming to sustain fisheries and ecosystems at their
current levels, that stocks should be increased and ecosystems restored (Pauly et al., 2002). Indeed this
approach has economic benefits where higher quotas could be sustained from a fishery with enhanced stock
abundance.
A number of principles, strategies and tools have been suggested to achieve successful ecosystem based
management of fisheries. One of the key principles that will need to be implemented is the precautionary
principle, which states that “In order to protect the environment, the precautionary approach shall be widely
applied by States according to their capabilities. Where there are threats of serious or irreversible damage,
lack of full scientific certainty should not be used as a reason for postponing cost-effective measures to
prevent environmental degradation” (United Nations, 1992). This principle requires that there be foresight
into the effects of fishing on the environment and the implications for future generations (Hall, 1999b). It
requires that potential environmental impacts are identified and that when there is uncertainty with respect to
the environmental impacts that priority should be given to conservation (Hall, 1999b).
To successfully manage marine ecosystems and fisheries a number of management approaches can be
implemented. Firstly, it has been widely recognised that to achieve sustainability in capture fisheries that
there needs to be a reduction in fishing effort (Hall, 1999b; Pauly et al. 2002). Secondly, government
subsidies are believed to encourage overcapitalization, inefficient harvesting, reduced potential economic
gains and ultimately have negative impacts on the sustainability of fisheries (Hilborn et al., 2003). Subsidies
allow overcapitalized fisheries to continue to operate and allow ‘profits’ to be generated even when
resources are overfished, promoting further exploitation of already overfished stocks (Pauly et al., 2003,
Pauly et al., 2002). Indeed, 17% of the landed value of fisheries in member countries of the Organisation for
Economic Co-operation and Development (OECD), of which Australia is a member, are subsidies (Hilborn
et al., 2003). The removal of subsidies would allow for a transparent evaluation of the economic profits from
fisheries. In particular, the removal of fuel subsidies would be expected to make long-distance journeys
uneconomical and therefore result in quasi marine reserves in areas that are long distances offshore (Pauly et
al., 2003).
18
One management option in ecosystem management is the use of marine protected areas (MPAs). Whilst
MPAs can offer varying degrees of protection, through regulation of gear types and season for example
(Hall, 1999b), the focus here is on no-take areas, or marine reserves. Marine reserves have been shown to be
an effective tool for rebuilding depleted stocks, for re-establishing trophic cascades that had collapsed due to
fishing pressure and even for enhancing fisheries in areas adjacent to the reserve (Mosqueira et al., 2000,
Roberts et al., 2001, Shears and Babcock, 2002). Illegal fishing in marine reserves is also relatively easy to
enforce with the use of satellite, compared with trying to enforce restrictions based on fishing gear or species
(Pauly et al., 2002). It should be noted, however, that MPAs will have minimal benefits for migratory species
with ranges that extend beyond the MPA (Pauly et al., 2002).
In terms of implementing a management plan and allocating controls to a fishery, there are a number of tools
evolving that can be used to predict the effects that removing numerous individuals of a population will have
on a community or ecosystem. In particular the Ecopath software package can be used to assess the
ecosystem effects of fishing as well as to assist in determining the effectiveness of management policy
options and placement of MPAs (Pauly et al., 2000b). Currently there are only Ecopath models available for
the Great Barrier Reef and Southern Tasmania in Australia, however, this could be a valuable management
tool for Australia’s fisheries into the future.
6.1 Management Perspectives in Australia
The observed trend of sequential harvesting in Australian fisheries is not unexpected given the lack of
management controls in the past. The expansion of the fishing industry in Australia was rapid and targeted a
number of different species (Kailola et al., 1993). Stocks were harvested in a technically advanced manner
by many fishers, where management controls and scientific information were limited (Kailola et al., 1993).
There were insufficient resources to study every species so scientific research was focussed on the valuable
species (Kailola et al., 1993). This allowed some valuable species, for example the western rock lobster, to
be well studied and well managed (Kailola et al., 1993). However a number of species were poorly studied
and managed, leading to overharvesting of these stocks.
Major improvements in the management of Australian fisheries occurred with the implementation of the
Fisheries Management Act 1991, which resulted in the formation of the Australian Fisheries Management
Authority and publication of yearly Fishery Status Reports for fisheries managed by the Australian
Government. Further action was then taken in 2005 with announcement of the $220 million Securing Our
Fishing Future package, which is designed to make Australian Government-managed fisheries sustainable
and profitable (Australian Government, 2005).
The Securing Our Fishing Future package addresses and utilises to some extent a number of the
management principles, strategies and tools that were described in order to successfully manage fisheries and
19
take into account ecosystem effects of fishing. Indeed the Ministerial Directive that was given with the
announcement of the Securing Our Fishing Future package stated that action was needed to ‘give overfished
stocks a chance to recover to an acceptable level in the near future’ and not just to sustain current levels. The
management strategy resulting from this directive is to be ecosystem based fisheries management that
incorporates the precautionary approach and uses management techniques such as the establishment of
MPAs and reduction in fishing effort. Despite this, the management of Australian fisheries still has a number
of limitations. The benefits and limitations of fisheries management of Australian Government-managed
fisheries proposed under the Securing Our Fishing Future initiative will be further discussed.
Although the Australian Fisheries Management Authority has stated that they are adopting ecosystem based
fisheries management in response to the Ministerial Directive, the management of Australian fisheries is still
largely based on reference points from single species management. Such reference points include stock
biomass, fishing mortality and maximum sustainable yield. These reference points will be used to allocate
quotas, where fisheries are to be managed using an individual transferable quota (ITQ) system. The transfer
of fisheries to ITQ management is expected to have a number of advantages, as the use of ITQs changes the
way a resource is viewed from a common to instilling a sense of ownership of the resource. Another
advantage of ITQ management is that it improves efficiency, where there is no incentive to acquire more
vessels or larger vessels than what is required to obtain the quota (Hannesson, 1991).
Another management approach being implemented in order to enhance the sustainability of Australian
Government-managed fisheries is to reduce effort by reducing the number of fishers competing for
resources. This approach is also aimed at maximising the profitability of fishing businesses. This was done
through a Business Exit Assistance program (Australian Government, 2005) where the Australian
Government purchased statutory fishing rights from fishers in a number of Commonwealth fisheries. Whilst
a reduction in effort in the fisheries from which the statutory fishing rights were purchased would be
expected to improve management of those fisheries, there are a number of problems that can be associated
with subsidies such as these. This is because it is the right to use a particular boat or fishing gear that is being
purchased, not the actual boat or gear itself. This may lead to the vessel or gear being used in other fisheries
and the subsidy may even be put toward further vessel modernizations for enhanced efficiency within other
fisheries (Pauly et al., 2002). Also, if buyback schemes are used regularly then fishers have the expectation
that there will be future buyback schemes and the potential risk of entering a fishery is lowered and this leads
to increased capacity within the fishery (Gooday, 2002). Furthermore, the fishing industry in Australia is
further subsidised for management services and intermediate inputs, such as diesel fuel.
Under the Draft Commonwealth Harvest Strategy (Australian Government, 2007), that has resulted from the
Securing Our Fishing Future initiative, it has been stated that ecosystem interactions need to be considered.
It states that where the targeted species is considered a keystone species the biomass reference points should
be increased. It has not been defined how the ecosystem interactions of target species will be determined or
20
to what level biomass reference points should be increased. The available scientific information on species
interactions and ecosystem effects of removing target species from Australian fisheries is limited.
Furthermore, of the $220 million package, only $6 million dollars has been allocated for additional science,
compliance and data collection, where this money is to be used for activities such as the establishment of
independent surveys, the development of electronic-licensing and the deployment of onboard cameras. How
species interactions are to be researched, or indeed whether they are to be researched, and where the funding
is to come from for that research, was not clearly stated in the Securing Our Fishing Future package.
Therefore the extent that species interactions can be accounted for under this initiative remains questionable.
The current biomass reference points for reducing exploitation rates and ceasing exploitation are 40% and
20% of prefished levels respectively (Hindmarsh and Talbot, 2007). Again there are questions as to how
these levels are to be assessed, particularly with respect to what exactly constitutes prefished levels. A
common problem in assessing prefished levels is the shifting baseline syndrome, where the assessed
prefished level may be, for example, the level when fisheries records were first kept, not the true levels of
former natural abundance. An example of the shifting baseline syndrome can be seen with dugong
populations in Moreton Bay. Dugongs in Moreton Bay were first harvested extensively by Aboriginal people
and then by early European colonists, where the levels prior to Aboriginal harvesting are not known. By the
late nineteenth century there were still reports of herds of dugongs with tens of thousands of dugongs in each
herd (Jackson et al., 2001). In the twentieth century there was widespread colonial exploitation of dugongs
before the fishery collapsed. Then in the 1970s a “large population” of dugongs were discovered in Moreton
Bay (Heinsohn et al., 1978). This “large” population was only 300 individuals (Jackson et al., 2001).
Even if prefished levels can be accurately determined, the biomass reference points of 40% and 20% are also
concerning as to the implications for species interactions. It would be expected if a population is at only one
or two fifths of its prefished level, that there would be consequences for other species which interact with
this species, for example through predator-prey relationships, but also on the entire ecosystem through
trophic cascades and changes in habitat complexity should the targeted species be an ecosystem engineer.
Whilst there is no scientifically prescribed level that a population should be retained at in order to maintain
its functioning within an ecosystem, perhaps the precautionary approach needs to be applied here.
In Australia, fishing groups are very vocal and politically active, even with their own political parties “The
Fishing Party” and “The Australian Fishing and Lifestyle Party” (The Australian Fishing and Lifestyle Party,
2007a; The Fishing Party, 2007). As such, the main issues associated with applying the precautionary
approach in Australia are political, where any moves to restrict fishing, particularly when there is scientific
uncertainty, are met with objections from large numbers of professional and recreational fishers. Current
examples of such situations include an appeal by RECFISH, Australia’s peak body representing recreational
fisher’s interests, against marine sanctuaries to protect grey nurse sharks due to uncertainty in grey nurse
shark population estimates (RECFISH Australia, 2007). Also, claims by the Australian Fishing and Lifestyle
21
Party that an increase in no-take areas in the Moreton Bay Marine Park (which currently has less than 1%
no-take areas) is unwarranted because of a lack of “any irrefutable scientific evidence to suggest that no-take
zones will improve biodiversity and boost fish stocks” (The Australian Fishing and Lifestyle Party, 2007b).
These examples demonstrate that where there is scientific uncertainty, that the political influence of fishing
groups may inhibit the application of the precautionary principle. It is my belief that the only way to resolve
this conflict will come through public education on the status and sustainability of our fisheries. Indeed,
fishers and conservationists both want the same thing – for there to be plentiful fish in the ocean. Maybe if
more fishers realised the benefits that management techniques, such as MPAs, can have for enhancing fish
stocks they would not be so strongly opposed to the idea.
6.2
Future Perspective for the Management of Australian
Fisheries
Whilst the recent introduction of the Securing Our Fishing Future initiative may be beneficial in the
management of Australian Government-managed fisheries, there are still further improvements needed in the
management of these fisheries. These improvements are particularly relevant if Australia is to phase out
destructive fishing practices, such as sequential depletion of fish stocks, by 2012 as stated in the World
Summit on Sustainable Development. Furthermore, the management of State-managed fisheries is currently
variable and could also be improved. Firstly, there needs to be a reduction in effort in fisheries and subsidies
should not be offered where this will lead to further environmental harm. Secondly, the aim should be to
increase stocks and not just sustain current levels. In assessing the level of a stock which is to be maintained
all efforts should be made to avoid the shifting baseline syndrome. This will have economic advantages and,
in the case where marine reserves are used, may also be able to restore trophic cascades and enhance
ecosystem functioning. Thirdly, ecosystem based management should be applied across all fisheries to
complement the existing single species management, and the ecosystem effects of removing species should
be evaluated and incorporated into management decisions. This will require enhanced scientific
understanding of species interactions and scientific research will need to occur for this to be achieved. Our
current knowledge of species interactions and ecosystem effects of fishing is limited and the precautionary
approach should be applied. Public education and awareness programs are needed to overcome the political
problems associated with management decisions.
22
7 CONCLUSIONS
The global trend of fishing down the food web was not observed in mean trophic levels in Australia. The
mean trophic level of Australian fisheries was found to be highly variable but generally increased during the
1970s, 1980s and 1990s. The variability in mean trophic level was a result of sequential depletion of stocks.
Sequential harvesting has occurred at a number of scales, from all fisheries within Australia, to within
multispecies fisheries, to within single species fisheries. This trend of sequential depletion is unsustainable
and is likely to have ecosystem scale consequences. These ecosystem consequences can occur when target
species of any trophic level are depleted and the effects may not be apparent for decades or even centuries to
come. Species interactions and the trophic effects of harvesting target species need to be considered in
ecosystem based fisheries management. Currently it has been stated that these interactions will be taken into
account in the management of Australian Government-managed fisheries, however the scientific data on
species interactions is limiting. These limitations need to be addressed if Australia’s fisheries are to be
sustainable and profitable into the future and if the destructive practice of sequential depletion is to be
phased out by 2012.
23
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