e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
available at www.sciencedirect.com
journal homepage: www.elsevier.com/locate/ecolmodel
Comparing trophic flows and fishing impacts of a NW
Mediterranean ecosystem with coastal upwelling systems
by means of standardized models and indicators
Marta Coll a,∗ , Lynne J. Shannon b , Coleen L. Moloney c ,
Isabel Palomera a , Sergi Tudela d
a
Institute of Marine Science (CMIMA-CSIC), Passeig Marı́tim de la Barceloneta 37-49, 08003 Barcelona, Spain
Marine and Coastal Management, Private Bag X2, Rogge Bay 8012, Cape Town, South Africa
c Marine Biology Research Institute, Zoology Department, University of Cape Town, Rondebosch 7701, Cape Town, South Africa
d WWF Mediterranean Programme Office, Canuda 37, 08002 Barcelona, Spain
b
a r t i c l e
i n f o
a b s t r a c t
Article history:
The NW Mediterranean has a number of structural features in common with upwelling
Received 1 April 2005
ecosystems. Therefore, an ecological model representing a NW Mediterranean exploited
Received in revised form 6 April
ecosystem was standardized and compared with four previously standardized models
2006
from coastal upwelling ecosystems: the Northern and Southern Humboldt (Chile and Peru
Accepted 12 April 2006
upwelling systems) and the Northern and Southern Benguela (Namibia and South Africa
Published on line 13 June 2006
upwelling systems). Results from biomasses, flows and trophic levels indicated important
differences between ecosystems, mainly caused by differences in primary production, which
Keywords:
was smallest in the NW Mediterranean Sea. However, principal component analysis (PCA)
Mediterranean
of biomasses and flows suggested a similar pattern between the NW Mediterranean and the
Upwelling ecosystems
South African systems due to the inclusion of an important fraction of the continental shelf
Ecological modelling
in both ecological models representing these areas. At the same time, diets of commercial
Trophic flows
species from the NW Mediterranean were more similar to Benguela than Humboldt species.
However, the relatively heavy fishing pressure in the NW Mediterranean ecosystem was
Ecosystem indicators
Fishing impact
highlighted relative to its primary production, and was evident from the large catches and
small primary production, largest flows from TL 1 required to sustain the fishery (%PPR),
the low trophic level of the catch (TLc ), high exploitation rates (F/Z), largest values in the
trophic spectra portraying catch: biomass ratio, the FIB index and the demersal: total catch
ration. Comparisons of %PPR, the trophic level of the community (TLco ), the biomass of
consumers and F/Z ratios seemed to capture the ecosystem effects of fishing: large in the
NW Mediterranean, Namibia and Peru upwelling systems. Small pelagic fish were the most
important component of the fisheries in the NW Mediterranean and Peruvian systems.
However, the smaller production and biomass ratios from the NW Mediterranean could
be an indirect indicator of intense fishing pressure on small pelagic fish, also in line with
results from consumption of small pelagic fish by the fishery, F/Z ratios and trophic spectra.
Moreover, similarities between the NW Mediterranean and Namibian systems were found,
mainly related to the demersal: total catch ratios, the FIB index, the relevance of gelatinous
zooplankton in the consumption of production and the importance of pelagic-demersal
∗
Corresponding author. Tel.: +34 93 230 95 43.
E-mail address: [email protected] (M. Coll).
0304-3800/$ – see front matter © 2006 Elsevier B.V. All rights reserved.
doi:10.1016/j.ecolmodel.2006.04.009
54
e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
coupling, in remarkable contrast to the other ecosystems. These similarities should be interpreted in terms of dynamic trajectories that the Namibian system has shown due to the
collapse of its pelagic ecosystem, partly due to fishing intensity, and the signs that the NW
Mediterranean could follow suit in the future.
© 2006 Elsevier B.V. All rights reserved.
1.
Introduction
1.1.
The NW Mediterranean Sea ecosystem
The Mediterranean region has been inhabited for millennia
and human settlements have been spreading continuously
along its coastal areas (Margalef, 1985). As a consequence,
marine ecosystems of the Mediterranean have been altered
in many ways over the centuries (Bianchi and Morri, 2000;
Papaconstantinou and Farrugio, 2000). Fishing activity has
been proposed as the first major human disturbance to coastal
areas (Jackson et al., 2001) and evidence of fishing activity
going back to ancient times can be found throughout the
Mediterranean Sea (Bas et al., 1985). Moreover, the development of fishing technologies and overcapitalization in recent
decades, with an increasing demand for marine resources,
is placing intensive pressure on the exploited ecosystem.
The current assessment from the NW Mediterranean suggests that demersal stocks are fully exploited or overexploited,
whilst some pelagic stocks also show signs of overexploitation
(Farrugio et al., 1993; Papaconstantinou and Farrugio, 2000; Bas
et al., 2003; Lleonart and Maynou, 2003; Lleonart, 2005).
In order to describe the structure and functioning of a relatively productive exploited ecosystem from the NW Mediterranean and assess the ecosystem effects of fishing, an Ecopath mass-balanced model (Pauly et al., 2000; Christensen and
Walters, 2004) was constructed. The model represents the continental shelf and upper slope area associated with the Ebro
River Delta (South Catalan Sea, NW Mediterranean) (Coll et al.,
2005) in 1994, when official landings were at their highest level
since the 1970s. Results from the ecological model showed
that the ecosystem was dominated by the pelagic compartment, with which main flows and biomasses were associated.
Small pelagic fish, mainly European sardine (Sardina pilchardus)
and European anchovy (Engraulis encrasicolus), were identified
as important components of the ecosystem, dominating the
pelagic fraction in terms of biomasses and catches. A calibration process of the model with available time series of data
(Coll et al., 2006) suggested that sardine would be involved
in wasp-waist trophic control situations (as defined by Rice,
1995 and Cury et al., 2000). Moreover, European hake (Merluccius merluccius) and medium-sized pelagic fish (mainly horse
mackerel Trachurus spp. and mackerel Scomber spp.) were also
important in terms of biomasses and trophic interactions. The
model also showed that the NW Mediterranean ecosystem
was highly impacted by fishing activity.
The studied NW Mediterranean ecosystem had a number
of structural features in common with upwelling ecosystems,
more so than with other known ecosystems that have been
modelled (e.g. Christensen and Pauly, 1993; Jarre-Teichman,
1998; Cury et al., 2000; Shannon et al., 2003; Heymans et
al., 2004; Sánchez and Olaso, 2004). These features include
the dominance of the pelagic compartment, the importance
of small pelagic fish in terms of catch and biomass and
their implication for wasp-waist flow control situations, the
key role of other pelagic fish, such as horse mackerel, the
importance of hake and the low development stage of the
ecosystem sensu Odum (1969) (Christensen, 1995). Moreover,
oceanographic conditions and local upwelling events in the
NW Mediterranean, mainly related to wind conditions, vertical mixing and stratification of water, fresh water inputs,
shelf-slope exchanges and density fronts (Estrada, 1996; Salat,
1996; Agostini and Bakun, 2002), greatly influence the productivity and fishing activity in the area. Nutrient enrichment
and relatively high concentrations of small pelagic fish occur
(Palomera, 1992; Estrada, 1996; Sabatés and Olivar, 1996; Salat,
1996; Lloret et al., 2004).
1.2.
Comparing ecosystem models
The Ecopath with Ecosim approach (EwE) has been widely
used to quantitatively improve the knowledge on structure and functioning of different marine ecosystems and, by
analysing ecological indicators provided directly from these
models, it has been possible to contextualize the fishing
impact and quantify its ecosystem effects (e.g. Christensen
and Pauly, 1993; Christensen, 1995; Roux and Shannon, 2004;
Sánchez and Olaso, 2004). In addition, this methodology has
been intensively applied to upwelling regions, enabling the
improvement of descriptions of ecosystem functioning and of
the importance of fishing activities and environmental factors in ecosystem dynamics (e.g. Jarre-Teichmann et al., 1998;
Heymans et al., 2004; Shannon and Cury, 2003; Neira and
Arancibia, 2004; Shannon et al., 2004a,b).
Among the analyses of exploited ecosystems undertaken
using the EwE approach, the comparison of ecological models
representing different situations of a given ecosystem through
time has been shown to be a useful exercise (e.g. Libralato et
al., 2002; Shannon et al., 2003; Heymans et al., 2004; Neira et
al., 2004). Furthermore, by standardizing models of different
ecosystems to achieve a common structure separating biological features from modelling artefacts, these comparisons
have been successfully applied to four different upwelling
ecosystems of the Humboldt and Benguela. These models represented different areas and periods (Moloney et al., 2005):
the Southern Humboldt upwelling ecosystem (Chilean system), the Northern Humboldt upwelling ecosystem (Peruvian
system), the Southern Benguela upwelling ecosystem (South
African system) and the Northern Benguela upwelling ecosystem (Namibian system).
In this study, and because of the similarities found between
the NW Mediterranean and upwelling ecosystems, the available ecological model from the South Catalan Sea, NW
Mediterranean, was reformulated to conform to the same
55
e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
standardized model as used for the upwelling ecosystems
and compared with those models presented in Moloney et al.
(2005). This comparison is especially relevant in the case of the
Mediterranean Sea, where there are few available time series
of data and where ecological modelling of exploited ecosystems it is still scarce (e.g. Libralato et al., 2002; Pinnegar and
Polunin, 2004; Coll et al., 2005).
The principal aims of the comparison were: (a) to assess
differences and similarities in the structure and functioning
of the NW Mediterranean ecosystem and the four upwelling
systems related to their intrinsic features and exploitation history; (b) to analyse the trophic information and ecological roles
of common important species in terms of biomasses, trophic
information and consumption; (c) to compare a selected group
of ecological indicators to assess differences in the ecosystem
impacts of fishing.
Taking into account the exploitation history of the five
ecosystems, selected indicators were analysed to test the
initial hypothesis that the NW Mediterranean model would
be ranked within the high impacted areas of the Namibian
and Peruvian systems, which were differentiated in terms of
ecosystem impact from the moderately impacted regions of
the Chilean and South African systems (Moloney et al., 2005).
2.
Methods
2.1.
Ecological models and standardization process
The standardized upwelling models in Moloney et al. (2005)
from the Humboldt and Benguela ecosystems represented
the Chilean system in 1992 (Neira and Arancibia, 2004),
the Peruvian system in 1973–1981 (Jarre-Teichmann et al.,
1998), the South African system in 1980–1989 (Shannon et
al., 2003) and the Namibian system in 1995–2000 (Roux and
Shannon, 2004). These models represent exploited ecosystems from where information on temporal dynamics of
their marine resources is available (e.g. Neira et al., 2004;
Heymans, 2004; Heymans et al., 2004; Shannon et al., 2003,
2004a,b).
The model of the Chilean system (1992) represented the
ecosystem under a moderate ENSO event, where main fish
stocks were described as not fully exploited and the ecosystem
was dominated by medium-sized pelagic fish (mainly Pacific
jack mackerel Trachurus symmetricus and hoki Macruronus
magellanicus), with medium to high abundance of smallsized pelagic fish (common sardine Strangomera bentincki and
anchovy Engraulis ringens) and Chilean hake (Merluccius gayi)
(Neira and Arancibia, 2004). The model of the Peruvian system (1973–1981) represented the period following the collapse
of the anchoveta (E. ringens) fishery, when sardine (Sardinops
sagax) biomass was increasing and no major ENSO events
were recorded (Jarre-Teichmann et al., 1998). The model of the
South African system (1980–1989) represented a period when
the ecosystem was dominated by Cape anchovy (E. encrasicolus) with reduced biomass of sardine (S. sagax), and when
stocks of round herring or redeye (Etrumeus whiteheadi), horse
mackerel (Trachurus trachurus capensis) and Cape hakes (Merluccius capensis and M. paradoxus) were believed to be healthy,
whilst the pelagic resources were well utilized (Shannon et al.,
2003). The model of the Namibian system (1995–2000) represented a period in which pelagic gobies (Sufflogobius bibarba-
Table 1 – Input and output parameters of the standardized Ecopath model for the NW Mediterranean ecosystem (1994)
Functional groups
Phytoplankton
Microzooplankton
Mesozooplankton
Macrozooplankton
Gelatinous zooplankton
Macrobenthos
European anchovy
Special small pelagic
Benthopelagic fishes
Cephalopods
Other small pelagics
Horse mackerel
Characteristic large pelagics
Tunas and swordfish
Juvenile hake
Adult hake
Demersal benthic feeders
Demersal pelagic feeders
Demersal chondrichthyans
Seabirds
Marine turtles
Cetaceans
Detritus + discards
B
P/B
Q/B
Catch
EE
TL
10.20
2.10
7.79
0.54
0.39
24.71
2.64
3.58
0.22
0.41
0.92
1.55
0.61
0.40
0.04
0.35
0.73
1.37
0.06
0.002
0.03
0.39
70.38
37.91
24.18
20.87
20.41
25.00
1.55
1.33
1.50
1.37
2.18
0.52
0.39
0.46
0.37
1.30
0.60
1.33
0.67
0.42
4.60
0.15
0.04
–
73.24
48.85
50.94
50.48
8.53
13.91
8.86
9.03
15.63
7.39
5.13
4.88
3.52
7.37
2.52
6.66
5.86
5.43
71.58
2.54
4.73
–
–
–
–
–
–
0.22
0.94
2.83
0.07
0.27
0.01
0.02
0.05
0.05
0.02
0.21
0.44
0.22
0.01
0.0001
0.0006
0.002
–
0.94
0.95
0.76
0.91
0.12
0.36
0.96
0.97
0.97
0.97
0.98
0.30
0.51
0.34
0.98
0.98
0.98
0.97
0.90
0.18
0.12
0.10
0.77
1.00
2.05
2.15
2.84
2.90
2.01
3.15
3.06
3.55
3.61
3.09
3.27
3.62
4.16
3.36
4.16
3.14
3.52
3.75
2.95
2.54
4.03
1.00
B = biomass (t km−2 ); P/B = production/biomass ratio; Q/B = consumption/biomass ratio; catch (t km−2 year−1 ); EE = ecotrophic efficiency;
TL = trophic level.
56
Table 2 – Diet composition matrix of the standardized Ecopath model for the NW Mediterranean ecosystem (1994) (predators are located by columns, prey by rows)
Phytoplankton
Microzooplankton
Mesozooplankton
Macrozooplankton
Gelatinous zooplankton
Macrobenthos
European anchovy
Special small pelagic
Benthopelagic fishes
Cephalopods
Other small pelagics
Horse mackerel
Characteristic large pelagics
Tunas and swordfish
Juvenile hake
Adult hake
Demersal benthic feeders
Demersal pelagic feeders
Demersal chondrichthyans
Seabirds
Marine turtles
Cetaceans
Detritus + discards
Import
Total
3
4
5
0.700
0.050
0.650
0.090
0.050
0.150
0.050
0.600
0.050
0.100
0.050
0.650
6
0.001
0.001
0.001
7
8
1.000
0.076
0.050
0.874
9
10
11
0.052
0.010
0.926
0.066
0.248
0.611
12
13
14
15
16
0.010
0.687
0.165
0.010
0.269
0.618
0.001
0.011
0.006
0.050
0.006
0.075
0.479
0.264
0.176
0.002
0.045
0.001
0.012
0.114
0.018
0.049
0.049
0.005
0.003
0.005
0.022
0.002
0.002
0.092
0.198
0.020
0.215
0.009
0.063
0.002
0.001
0.751
0.141
0.001
0.015
0.566
0.53
17
0.027
0.006
0.038
0.844
0.002
0.001
0.005
0.010
0.001
0.066
0.059
0.037
0.013
0.004
0.153
0.128
0.039
0.013
18
19
20
0.406
0.142
0.002
0.450
0.044
0.038
0.006
0.424
0.104
0.063
0.152
0.003
0.011
21
22
0.373
0.013
0.294
0.077
0.156
0.003
0.067
0.002
0.001
0.036
0.001
0.010
0.074
0.059
0.001
0.001
0.012
0.030
0.021
0.093
0.005
0.032
0.010
0.250
0.210
0.150
0.150
0.992
0.002
0.010
0.002
0.400
1.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
0.371
0.312
0.293
0.400
0.406
1.0
1.0
1.0
e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
2
Table 3 – Comparative trophic regimes of small and medium-sized pelagic fish and hake species for the five models
Chile
Anchovy
Engraulis ringens:
Phytoplankton (TL = 2.1)
Sardine
South Africa
Namibia
Engraulis ringens:
Phytoplankton,
mesozooplankton (TL = 2.7)
Engraulis encrasicolus:
Mesozooplankton,
macrozooplankton
(TL = 3.54)
Sardinops sagax:
Phytoplankton, micro- and
mesozooplankton
(TL = 2.99)
Etrumeus whiteheadi:
Mesozooplankton,
macrozooplankton
(TL = 3.64)
Trachurus t. capensis:
Macrozooplankton,
mesozooplankton, small
pelagics (TL = 3.72)
Scomber scombrus:
Macrozooplankton,
mesopelagic fishes (TL = 3.8)
Engraulis encrasicolus:
Phytoplankton, meso- and
macrozooplankton (TL = 3.0)
Sardinops sagax:
Mesozooplankton,
phytoplankton (TL = 3.2)
Special small pelagic
Strangomera bentincki:
Phytoplankton (TL = 2.1)
Horse mackerel
Trachurus symmetricus:
Macrozooplankton (TL = 4)
Trachurus murphyi: Anchovy,
macrozooplankton (TL = 3.7)
Characteristhic large
pelagic
Macruronus magellanicus:
Macrozooplankton,
mesozooplankton (TL = 4.2)
Scomber japonicus:
Macrozooplankton,
anchovy (TL = 3.7)
Adult hake (1)
Merluccius gayi: Small
pelagics, juv. hake,
macrobenthos (TL = 3.5)
Merluccius gayi: Small
pelagics, Macrozooplankton
(TL = 4.0)
Adult hake (2)
Juvenile hake
Merluccius gayi: Small
pelagics, juv. hake (1),
macrobenthos (TL = 3.4)
Merluccius gayi:
Macrozooplankton, small
pelagics, mesozooplankton
(TL = 3.7)
Merluccius capensis:
Medium-sized pelagics, juv.
hake (1–2), small pelagics,
macrozooplankton
(TL = 4.66)
Merluccius paradoxus:
mesopelagics,
macrozooplankton, juv.
hake (2), cephalopods
(TL = 4.49)
M. paradoxus and capensis:
Macrozooplankton,
mesopelagics, small
pelagics (TL = 3.9 – 4.0)
NW Mediterranean
Engraulis encrasicolus:
Mesozooplankton,
macrozooplankton (TL = 3.15)
Sardinops sagax:
Phytoplankton, meso- and
macrozooplankton (TL = 2.7)
Sufflogobius bibarbatus:
Macrozooplankton,
macrobenthos,
mesozooplankton (TL = 3.2)
Trachurus t. capensis:
Macrozooplankton,
mesozooplankton,
macrobenthos (TL = 3.6)
Merluccius capensis: Small
pelagics, medium-sized
pelagics, juv. hake (1),
macrozooplankton (TL = 4.5)
Sardina pilchardus:
Mesozooplankton,
phytoplankton (TL = 3.6)
Trachurus spp.:
Meso- and macrozooplankton,
macrobenthos, small pelagics
(TL = 3.27)
Scomber spp.:
Macro- and mesozooplankton,
small pelagics, macrobenthos
(TL = 3.55
Merluccius merluccius: Small
pelagics, mesopelagics,
demersal fishes (TL = 4.10)
e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
Peru
Merluccius paradoxus:
Macrozooplankton,
mesopelagics, small
pelagics (TL = 4.3)
Merluccius capensis: Small
pelagics, macrozo opl
ankton (TL = 4.03)
Merluccius merluccius:
Macrobenthos, mesopelagics,
small demersal fishes
(TL = 3.45)
57
58
e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
tus), jellyfish (mainly Chrysaora sp.) and horse mackerel (Trachurus t. capensis) dominated the ecosystem and sardine (S.
sagax) and anchovy (E. encrasicolus) were at low biomass levels
after the collapse of pelagic fisheries (1960–1970s) (Roux and
Shannon, 2004).
The ecological model representing the 1994 annual situation of the continental shelf and upper slope ecosystem
of the South Catalan Sea, NW Mediterranean (Coll et al.,
2005) was standardized following the generic model structure proposed by Moloney et al. (2005). The standardization process required aggregation of the initial 40 functional groups and splitting of the initially combined microand mesozooplankton group into two groups according to
data from Calbet et al. (2001). New data of P/B and Q/B
ratios for zooplankton were adapted from Sánchez and Olaso
(2004), whilst trophic information was modified to account
for the split zooplankton groups (Demirhindi, 1961; Bell and
Harmelin-Vivien, 1983; Ben Salem, 1988; Tudela and Palomera,
1997; Stergiou and Karpouzi, 2002). The resulting functional
groups and the input parameters for the model are listed in
Tables 1 and 2.
The European sardine was allocated to the “special small
pelagic” group following the comparison criteria, the “other
small pelagic fish” group included the bogue (Boops boops) and
the sardinella (Sardinella aurita), whilst the “characteristic large
pelagic” group included the mackerel (Scomber scombrus and S.
japonicus). Unlike the other four ecosystems previously examined, three species were included within the horse mackerel
group: Trachurus trachurus, T. mediterraneus and T. picturatus.
Atlantic bonito (Sarda sarda), swordfish (Xiphias gladius) and
bluefin tuna (Thunnus thynnus) were included in the “tunas and
swordfish” group.
Fig. 1 – Integrated biomass and flows for the five models: (a) total biomass, excluding detritus (t km−2 ); (b) total biomass,
excluding detritus and all plankton groups (t km−2 ); (c) primary producer standing stock (t km−2 ); (d) total production,
excluding all plankton groups (t km−2 year−1 ); (e) total consumption, excluding all plankton groups (t km−2 year−1 ); (f) total
catches (t km−2 year−1 ). Values from the NW Mediterranean are highlighted.
e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
59
Table 4 – Eigenvalues and cumulative percent variance
explained by axes of the principal component analysis
(PCA) from Figs. 1, 7, 11 and 13
Axes
1
2
3
4
PCA1 (Fig. 1)
Eigenvalues proportion
Cumulative % variance
0.62
62.5
0.31
93.8
0.06
99.7
0.003
100.0
PCA2 (Fig. 7)
Eigenvalues proportion
Cumulative % variance
0.72
71.8
0.19
90.6
0.07
97.5
0.03
100.0
PCA3 (Fig. 11)
Eigenvalues proportion
Cumulative % variance
0.42
41.7
0.29
70.6
0.25
95.2
0.05
100.0
PCA4 (Fig. 12)
Eigenvalues proportion
Cumulative % variance
0.33
33.05
0.28
61.5
0.22
83.4
0.17
100.0
Fig. 3 – Discrete trophic level spectra of biomass (t km−2 ) for
the five models.
2.2.
Comparison of trophic information and ecological
indicators
After standardization, diet information of small and medium
sized pelagic fish and juvenile and adult hake were compared across ecosystems. The trophic behaviour of the European anchovy and European hake is well known from the NW
Mediterranean Sea (Tudela and Palomera, 1997; Bozzano et
al., 1997, 2005), however the information on other small and
medium-sized pelagic fish from this area is qualitative and
outdated (Massutı́ and Oliver, 1948; Andreu and RodriguezRoda, 1951; Demirhindi, 1961; Bell and Harmelin-Vivien, 1983;
Ben Salem, 1988). Therefore, the quantitative diet information used for the comparison was that from the mass-balance
model for the NW Mediterranean Sea (Coll et al., 2005).
For various ecological groups, the trophic level (TL),
which identifies the position of organisms in the food chain
(Lindeman, 1942; Odum and Heald, 1975), was also analysed.
By convention, primary producers and detritus have TL = 1;
values for other groups are determined using mass-balance
models, gut content analysis or isotope data (Stergiou and
Karpouzi, 2002). The TL can be formulated as following:
TLj = 1 +
n
DCji TLi
i=1
Fig. 2 – Principal component analysis applied to integrated
biomass and flows for the five models (Fig. 1). Eigenvalues
and percent variance explained by axes are shown in
Table 4.
where j is the predator of prey i, DCji the fraction of prey i in
the diet of predator j and TLi is the trophic level of prey i.
In addition, a similar procedure to that followed by Moloney
et al. (2005) was applied to compare various population and
ecosystem indicators derived from the NW Mediterranean
model with the four previously standardized models from
upwelling ecosystems (Rochet and Trenkel, 2003; Christensen
and Walters, 2004; Cury et al., 2005; Moloney et al., 2005).
Integrated biomasses and flows included in the comparison
were total biomass (t km−2 ), excluding detritus and excluding detritus and plankton groups to avoid assessment problems, primary production (t km−2 year−1 ), total production
(t km−2 year−1 ), total consumption (t km−2 year−1 ) and total
catches (t km−2 year−1 ). Moreover, ratios of biomass, production and catch of small pelagics: large hake and large pelagics
(Bsp /Blp , Psp /Plp , Csp /Clp ), of planktivores: piscivores (Bpl /Bpc ,
Ppl /Ppc , Cpl /Cpc ) and of planktivores: total consumers (Bpl /Bt ,
Ppl /Pt , Cpl /Ct ) were analysed. At the same time, consumption
of total production (t km−2 year−1 ) by predator groups excluding zooplankton and benthos, consumption of small pelagic
fish production (t km−2 year−1 ) by their predators (including
the fishery) and total biomass (t km−2 ) per integer trophic level
were also included in the comparison. For an overall interpretation of these results, principal component analyses (PCA)
60
e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
were performed on these data. Input data was centered and
scaled by columns to give similar importance to all biomasses
and flows. In addition, the trophic spectra for biomass, catch
and the catch: biomass ratio following Gascuel et al. (2005)
Fig. 5 – Trophic level of the community TLco (excluding
TL = 1) and total catches TLc .
Fig. 6 – Flows from TL 1 required to sustain the catches for
the five models in terms of flows (t km−2 year−1 ) and in
percentage. Data of the NW Mediterranean is highlighted.
Fig. 4 – Trophic spectra of (a) biomass; (b) catches; (c) catch:
biomass ratios for the five models excluding the gelatinous
zooplankton and macrobenthos groups.
was also analysed and compared between the five ecosystem
models.
The average trophic level of the catch (TLc ) and the average trophic level of the community (TLco ) excluding TL = 1
were also included. The former reflects the strategy of a fishery in terms of food web components selected and is calculated as the weighted average of the TL of harvested species
(Pauly et al., 1998; Christensen and Walters, 2004), whilst the
later reflects the structure of the community and is calculated as the weighted average of the TL of all the species
within the ecosystem (Rochet and Trenkel, 2003). Both indicators have been shown to decrease when fishing impact
increases because large predators are removed from ecosystems whilst lower trophic level organisms prevail (Pauly et al.,
1998; Jennings et al., 2002; Pauly and Palomares, 2005).
In order to compare the ecological footprint of fishing
activities, the primary production and detritus (flows from
TL 1) required to sustain fisheries (PPR; typically expressed
as t km−2 year−1 ) was included in the comparison. The PPR is
obtained by back calculating the flows, expressed in primary
61
e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
production and detritus equivalents, for all pathways from the
caught species down to the primary producers and detritus
and increases with fishing intensity (Pauly and Christensen,
1995; Christensen and Walters, 2004). The PPR can be formulated as:
PPR =
paths
⎡
⎣ Yi ×
Pi
j,i
of the NW Mediterranean was obtained from the Institute of
Marine Science (CMIMA-CSIC, Barcelona, Spain) through the
fishermen associations of the area and from the regional government of Catalonia.
The FIB index is formulated as following:
⎤
Qj
Pj × EEj
× DCj,i ⎦
where Yi is the catch of a given group i, P the production of
predator j, Q the consumption of predator j, DC the diet composition of each predator j/prey i interaction in each path and
EE is the ecotrophic efficiency, or the proportion of the production that it is used within the system due to consumption
or is exported from the system (e.g. in terms of catches). This
index can also be expressed relative to the primary production and detritus of the ecosystem (%PPR). At the same time,
the exploitation rates (F/Z, fishing mortality to total mortality)
by ecological group were also considered. These indexs also
increase with fishing (Pauly and Christensen, 1995; Rochet and
Trenkel, 2003).
Finally, the demersal: total catch ratio, the Fishing in Balance (FIB) index (Christensen, 2000; Pauly et al., 2000) and the
plot of TLc and total catch (Pauly et al., 1998) were calculated
for the NW Mediterranean case study from an available time
series of catches (from 1976 to 2003) and compared results
from Benguela ecosystems (Cury et al., 2005). The catch series
FIB = log
Y 10TLi
i ik TL
Y 10
i i0
i
where Yik is the catch of species i during the year k, Yi0 the
catch of species i during the year at the start of a time series
and which serves as an anchor and TLi is the trophic level
of species i. Values of FIB = 0 indicate that a decrease in the
trophic level of the catch is matched by an increase in catch
because of higher production at low trophic levels. In contrast,
when the FIB index increases (>0) this indicates that there is
an expansion of the fishery or bottom-up effects occur. The
FIB index decreases (<0) when discarding occurs and is not
considered in the analysis or when the fisheries impact on
the ecosystem is so high that its functioning is impaired (Pauly
and Watson, 2005).
The demersal: total catch ratio informs of the origin of
catches (whether they come from the demersal or the pelagic
habitat) and it can increase with fishing intensity (Rochet
and Trenkel, 2003), although it has also been related to nutrient availability and eutrophication (De Leiva Moreno et al.,
2000).
Fig. 7 – Ratios of small pelagic fish: large hake and large pelagic fish, of planktivores: piscivores and of planktivores: total
consumers (excluding plankton, macrobenthos and detritus) for catches (a–c), biomass (d–f) and production (g–i) for the five
models. Values for the NW Mediterranean are highlighted.
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e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
3.
Results
3.1.
Trophic information on key species
The principal prey groups of small and medium-sized pelagic
fish and hake species from the different ecosystems are summarized in Table 3. European anchovy feed mainly on mesozooplankton, being trophically more similar to Cape anchovy
in the South African system, and to a lesser extent in the
Namibian system, than to anchoveta in the Humboldt ecosystem, which feed mainly on phytoplankton. This was also in
line with the trophic levels of anchovy in the different ecosystems (Table 3). The European sardine, feeding on micro- and
mesozooplankton, and to lesser extent on phytoplankton, was
trophically similar to sardine and round herring in the South
African system and to sardine in the Peruvian systems, whilst
sardine in the Namibian system and common sardine in the
Chilean one were considered to be more dependent on phytoplankton.
Medium-sized pelagic fish in the NW Mediterranean Sea
showed some trophic differences from upwelling ecosystems,
where the TL of horse mackerel was the lowest (Table 3). Differences in trophic data of horse mackerel (Trachurus spp.)
Fig. 9 – Exploitation rates (F/Z) of (a) anchovy, Sardinops and
the special small pelagic and (b) adult hake and juvenile
hake for the five models (see Table 3 for species scientific
name by ecosystem).
Fig. 8 – Principal component analysis applied to catch,
biomass and production rations of small pelagic fish: large
hake and large pelagic fish, of planktivores: piscivores and
of planktivores: total consumers (excluding plankton,
macrobenthos and detritus) for the five models (Fig. 7).
Eigenvalues and percent variance explained by axes are
shown in Table 4.
were mainly due to the higher intake of mesozooplankton
and the importance of macrobenthos in the NW Mediterranean diet, also observed to a lesser degree in the Namibian system, whereas the diet of horse mackerel in the other
ecosystems examined consisted mainly of macrozooplankton. The trophic behaviour of mackerel (Scomber spp.) from the
NW Mediterranean also showed some differences due to the
occurrence of macrobenthos in the diet of the Mediterranean
group.
The diet of adult European hake was similar to hake in
the Namibian system and to a lesser extent in the Chilean
one; however, lower rates of cannibalism were displayed. The
trophic behaviour of juvenile European hake was different
from that in the upwelling systems, with some similarities
to the Chilean species, because its main prey were macrobenthos and mesopelagic fish. Zooplankton and small pelagic fish
were more important in the diet of juvenile hake species in the
other systems. Juvenile hake in upwelling areas are mainly
pelagic-feeders and show high rates of cannibalism by adult
hake (e.g. Payne et al., 1987; Punt and Leslie, 1995). The juvenile
European hake in the NW Mediterranean seems to undertake nocturnal vertical movements through the water column
to prey on mesopelagic fish, whilst it preys on macrobenthos and small demersal fish during the day (Larrañeta, 1970;
Bozzano et al., 2005). The juvenile Cape hake have been found
e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
to remain high in the water column, possibly in order to avoid
cannibalism from adult hake (Pillar and Barange, 1995; Huse
et al., 1998). The vertical movements displayed by European
hake cannot be explained by cannibalism because of the lower
rates, although they could be related to the fact that juvenile
hake is escaping from other potential predators or could also
be related to differences in prey availability (Bozzano et al.,
2005). However, it could be argued that the lower incidences
of cannibalism that are observed are due to the vertical migration (successfully reducing cannibalism).
The TL of small pelagic fish and adult hake from the NW
Mediterranean fell between those for these species in the
Humboldt and Benguela ecosystems (Table 3). However, diet
information and TLs of small and medium sized pelagic fish
showed more similarities with the Benguela than Humboldt
ecosystems. Moreover, these results emphasized the significance of pelagic-demersal coupling in the NW Mediterranean
model, reflected by the importance of macrobenthos in the
diets of horse mackerel, mackerel and juvenile European hake.
A strong pelagic-demersal link has also been noted in the case
of the pelagic goby and horse mackerel in the Namibian system (Roux and Shannon, 2004).
Fig. 10 – (a) Fishing-in-balance and demersal: total catch
ratio (D/T); dotted lines indicate the seven fishing periods
discussed in the text and (b) plot of mean trophic level of
the catch and total catch for the NW Mediterranean case
study (1976–2003).
3.2.
63
Integrated biomasses, flows and trophic levels
The NW Mediterranean ecosystem was characterised by
smaller biomasses of consumers (t km−2 , both excluding
detritus, and excluding detritus and plankton groups) than
Humboldt and Benguela ecosystems (Fig. 1(a) and (b)). The
NW Mediterranean also had the lowest primary production
(t km−2 year−1 ), total production (t km−2 year−1 ) and total consumption (t km−2 year−1 ) that reflected the small dimensions
of the ecosystem in terms of flows per unit of area compared to upwelling ecosystems (Fig. 1(c–e)). However, the NW
Mediterranean showed the third highest catch (t km−2 year−1 )
of the comparison, higher than in the Benguela ecosystems (Fig. 1(f)). Results from principal component analysis
(PCA) applied to this data showed that the NW Mediterranean pattern was similar to that of the South African system, even though the latter system showed higher flows
and biomasses (Fig. 2(a) and (b)). The high biomass of consumers from the Namibian system (excluding detritus and
plankton groups) set this model apart from the rest, which
was related to the benthic biomass necessary to sustain the
trophic requirements of the pelagic goby and could be a model
artefact (Moloney et al., 2005). The Peruvian and Chilean
systems showed similar patterns but differ due to primary
production. Three factors explained 99.7% of the variance
(Table 4).
Analyzing the ecosystem models in terms of biomass
(t km−2 ) by discrete trophic levels (Fig. 3), it was also seen that
the NW Mediterranean had the smallest biomasses per discrete trophic level (TL), where biomass of TL II was higher
than TL I, as in the Namibian and Peruvian systems. The
trophic spectra by trophic levels II–V, excluding the gelatinous
zooplankton and macrobenthos groups, are shown in Fig. 4;
the NW Mediterranean biomass spectra being the smallest in
amplitude of all the cases examined (Fig. 4(a)). Moreover, it is
similar to Chilean system with the exception being that the
NW Mediterranean biomass spectrum lacked the marked second peak at TL 4 that resulted from characteristic large pelagics and large horse mackerel in the Chilean region. In addition,
some similarities were found when comparing the trophic
spectra of catches from the NW Mediterranean case study
and Benguela ecosystems, where the catch is mainly based
on organisms with TL 3 and TL 3.5, respectively (Fig. 4(b)).
It is important to highlight the similar shape of the trophic
spectrum of biomass and catch in the NW Mediterranean case
study (as in the Peruvian system). This is related to the multispecificity of the fishery, where fishing activity mainly targets
all that can be fished in the ecosystem (excluding plankton and benthic invertebrates) (Coll et al., 2005). The trophic
spectra of catch: biomass ratios for the NW Mediterranean
showed the highest values between TLs 3 and 3.7 (Fig. 4(c)).
The two distinguished peaks within this range highlight the
intense fishing pressure on organisms with TLs ≈ 3.1 (mainly
small pelagic fish) and TL ≈ 3.5 (mainly demersal fish and
cephalopods).
The mean trophic level of the community (TLco ) excluding TL = 1 ranged from 2.40 to 3.08 and was smallest in the
NW Mediterranean ecosystem, followed by the Namibian and
Peruvian systems (Fig. 5). The mean trophic level of the catch
(TLc ) for the NW Mediterranean was higher than the values
64
e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
obtained for Humboldt ecosystems and lower than those in
Benguela (Fig. 5).
Comparing the flows from TL = 1 required to sustain
the fishery (PPR) (Fig. 6) it was seen that the total flow
(t km−2 year−1 ) was the smallest in the NW Mediterranean,
whilst the percentage of the PPR (%PPR) was the highest (39%), followed by the Peruvian and the Namibian systems.
The importance of small pelagic fish in the NW Mediterranean and Peruvian systems can be observed from the analysis of the catch ratio of small pelagics: adult hake and
large pelagic fish (Csp /Clp ), of planktivores: piscivorous fish
(Cpl /Cpc ) and of planktivores: total consumers (excl. plankton,
macrobenthos and detritus) (Cpl /Ct ) (Fig. 7(a–c)). The biomass
ratio of small pelagics: adult hake and large pelagics (Bsp /Blp )
and of planktivores: piscivores fish (Bpl /Bpc ) was low in the
NW Mediterranean and similar to the Chilean and Namibian
systems (Fig. 7(d) and (e)). On the other hand, the biomass
ratio of planktivores: total consumers (Bpl /Bt ) was higher in
the NW Mediterranean than in the Peruvian and Namibian
systems, but lower than in the South African and Chilean
upwellings (Fig. 7(f)). The production ratio of small pelagics:
adult hake and large pelagic fish (Psp /Plp ) and of planktivores:
piscivores fish (Ppl /Ppc ) was also low in the NW Mediterranean
case study, similar to the Namibian and Chilean systems
(Fig. 7(g) and (h)), whilst the production ratio of planktivores:
total consumers (Ppl /Pt ) was higher than for the Benguela
ecosystems but lower than those in the Humboldt ecosystems (Fig. 7(i)). Results from PCA applied to these indicators
show that NW Mediterranean pattern is similar to that of the
South African and Chilean systems, and to a lesser extend
to the Namibian one, mainly due to similar values of Bpl /Bt
(Fig. 8(a) and (b)). Three factors explained 97.5% of the variance (Table 4).
Fig. 11 – Main partitioning (%) of total consumption of production by predators (excluding zooplankton and benthos) for the
five models (a–e). Total consumption is reflected in the relative sizes of the portion pies.
e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
3.3.
Fig. 12 – Principal component analysis applied to
partitioning (%) of total consumption of production by
predators (excluding zooplankton and benthos) for the five
models (Fig. 11). Eigenvalues and percent variance
explained by axes are shown in Table 4.
The exploitation rates (F/Z) (Fig. 9) had higher values in
the NW Mediterranean ecosystem, followed by the Peruvian,
Chilean and South African systems, and were zero for the
Namibian system in the case of the small pelagic fish due to
the collapse of the fisheries (Heymans et al., 2004; Roux and
Shannon, 2004).
The analysis of time series of demersal catches: total
catches, the FIB index and the plot of mean TLc and total
catches calculated for the NW Mediterranean from 1976 to
2003 (Fig. 10(a) and (b)) showed an increase of the demersal
fraction in the catch (from 22% to 46%) and of the mean TLc .
Fig. 10(a) and (b) can be interpreted in seven different periods
(dotted lines) corresponding with expansions of the fishery
and the intense exploitation of available resources, mainly of
the pelagic fraction. In this context, the FIB index decreased
from 1976 to 1978 and increased from 1978 to 1983. The value
decreased again from 1989 to 1990, increased from 1990 to
1994 (the year modelled in this study, and for which FIB was
at its maximum), and decreased from 1994 to 1998, during
which period small pelagic fish decreased in their contribution to total catches. From 1998 to 2000, there was a moderate
increase in the FIB index, followed again by a decrease till 2003,
where negative values have been shown for the last part of the
series.
65
Consumers of production
The largest consumers of production, after excluding zooplankton and benthic invertebrates, were the small pelagic
fish in all the case studies (Fig. 11). Results from PCA showed
that NW Mediterranean pattern are more similar to Benguela
ecosystems, mainly due to similar values of consumption by
mammals and turtles, cephalopods and demersal fish and
chondrichthyans (Fig. 12(a)). However, the NW Mediterranean
and Peruvian systems show similar patterns of consumption of production by small pelagic fish. In addition, the NW
Mediterranean shows similarities with the Namibian system
due to the relatively high proportion of consumption by jellyfish (Fig. 12(b)). Three factors explained 95.2% of the variance
(Table 4).
When analysing the impact of consumption of the production of small pelagic fish (Fig. 13), the fisheries were by far
the most important group in the NW Mediterranean, which
was set apart from the other ecosystems when a PCA was
performed (Fig. 14(a)). Three factors explained 83.36% of the
variance (Table 4). Moreover, consumption of production by
fisheries was also important in the Chilean and Peruvian systems, but had collapsed off the Namibian one by the late
1990s. Cephalopods, apex fish predators and other demersal
fish also played an important role in the NW Mediterranean
region, whilst cephalopods were also important in Benguela
ecosystems. As in the Peruvian system, marine mammals and
seabirds showed low impact in terms of consumption in the
NW Mediterranean, whilst the importance of these groups in
Benguela ecosystems, and to a lesser extent in the Chilean
system, was higher.
4.
Discussion and conclusions
Results from comparisons of biomasses, flows and trophic levels showed expected important differences between ecosystems resulting from differences in primary production, being
lower in the NW Mediterranean Sea, followed by the Benguela
and Humboldt ecosystems. This is in line with production
from the Mediterranean Sea (Bosc et al., 2004) and with differences in transfer efficiencies, which are higher in the NW
Mediterranean than in the other models included in the comparison (Jarre-Teichmann et al., 1998; Neira and Arancibia,
2004; Shannon et al., 2003; Heymans et al., 2004; Coll et al.,
2005) and indicates that the NW Mediterranean ecosystem
is food limited. This relates to previous observations suggesting that transfer efficiencies from primary and secondary
production decrease with increasing primary production so
that oligotrophic areas can be more efficient than highly productive ones (Cushing, 1975). Moreover, Alcaraz et al. (1985)
reported that the ratio between zooplankton and phytoplankton biomass was higher in the Western Mediterranean than
in the NW African upwelling regions, suggesting a relatively
high ecological efficiency in the Mediterranean Sea.
On the other hand, results from the trophic spectra analysis, total catches and trophic levels highlighted structural differences among Benguela, Humboldt and NW Mediterranean
food webs. This is in line with results from diets of commercial
species that reflect intrinsic ecosystem features. Differences
66
e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
Fig. 13 – Main partitioning (%) of total consumption of small pelagic fish production by their predators (including the fishery)
for the five models (a–e). Total consumption is reflected in the relative sizes of the portion pies.
in the diet of medium-sized pelagic fish may be also due to
limited availability of data for these species in the Mediterranean Sea (Stergiou and Karpouzi, 2002). In the case of horse
mackerel this may be due also to the fact that the diet for the
NW Mediterranean case study has been assumed applicable
to both small and large horse mackerel due to lack of detailed
information, whilst ontogenetic information is available for
the upwelling ecosystems.
Results from biomasses, flows and consumption of production also showed that NW Mediterranean had similar patterns to the South African system. This could be related to
the fact that both models from NW Mediterranean and South
African systems include a part of shelf habitat in comparison with the other ecosystems compared. The South African
model includes the Agulhas Bank, which is important in terms
of shelf habitat (Shannon and Jarre-Teichmann, 1998) and
the NW Mediterranean model includes the continental shelf
ecosystem associated with the Ebro River Delta (Coll et al.,
2005).
From the analysis of consumption of production it is also
shown that the role of benthopelagic fish in the NW Mediterranean was smaller than in the other models. This could be
related to the large proportion of continental shelf within
the NW Mediterranean model or to the general problems of
biomass estimation of these daily migratory species.
Results from this comparison also underlined the higher
impact of fishing within the NW Mediterranean ecosystem
relative to the primary production, reflecting the high fishing
pressure in that area. This was highlighted by high catches and
low primary production, low total biomass and low total sec-
e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
Fig. 14 – Principal component analysis applied to
partitioning (%) of total consumption of small pelagic fish
production by their predators (including the fishery) for the
five models (Fig. 13). Eigenvalues and percent variance
explained by axes are shown in Table 4.
ondary production, high %PPR, low TLc and large values for the
trophic spectra of catch: biomass ratio (Pauly and Christensen,
1995; Pauly et al., 1998; Rochet and Trenkel, 2003). In the case
of the NW Mediterranean, total biomasses and flows were also
lower and these indicators are expected to decrease with high
fishing pressure, although they can also be related to ecosystem size and might be difficult to predict due to indirect effects
of fishing through food webs (Rochet and Trenkel, 2003).
In addition, the NW Mediterranean showed high exploitation rates (F/Z for small pelagics and hake) and in the case of
European sardine the value was higher than 0.5, the limit reference point above which overexploitation is likely to occur
(Patterson, 1992; Mertz and Myers, 1998; Rochet and Trenkel,
2003) and similar to the F/Z ratios for sardine in the Peruvian system. Moreover, the exploitation rate for juvenile and
adult hake was the highest, and in the case of the adult hake
the value was higher than the recommended rate of 0.8 for
groundfish stocks (Mertz and Myers, 1998; Rochet and Trenkel,
2003).
Furthermore, the comparison of %PPR, TLco , total biomass
of consumers (excluding plankton and macrobenthos) and
exploitation rates (F/Z) seemed to capture the ecosystem
effects of fishing (Rochet and Trenkel, 2003), larger in the
Namibian, Peruvian and NW Mediterranean systems. The
analysis of biomass by discrete trophic level also showed similarities between these ecosystems.
67
Small pelagic fish were very important for the NW Mediterranean and Peruvian fisheries, with high catch of small pelagics: adult hake and large pelagics, of planktivores: total consumers and, to a lesser extent, of planktivores: piscivores fish.
This was also highlighted with the size spectra and the low
TLc . However, important differences in the production and
biomass of small pelagic fish were seen between these two
areas; the ratios from the NW Mediterranean were more similar to the Benguela and Chilean systems than to the Peruvian
one. This could be an indirect indicator of the intense fishing
pressure of small pelagic fish in the NW Mediterranean, also in
line with results for consumption of small pelagic fish production and exploitation rates (F/Z). Therefore, taking into account
the differences in ecosystem production, present results suggest that marine resources in the NW Mediterranean ecosystem have been subjected to high fishing pressure in line with
Coll et al. (2005).
The FIB index and the demersal: total catch ratio applied
to the NW Mediterranean showed various periods of expansion of the fishery (increasing values of FIB), but also of intense
impact of the exploited pelagic food web (decreasing values of
FIB). The expansion periods of the fishery most likely resulted
from governmental aids to the fishing sector and the implementation of technological advances (Bas et al., 1985; Farrugio
et al., 1993; Papaconstantinou and Farrugio, 2000). Due to the
fact that demersal: total catch increased, an eutrophication
process in the area cannot be identified (De Leiva Moreno et
al., 2000). Time periods where the FIB index was decreasing
could be related to periods with high impact of fishing and
decreased stock sizes of targeted species (mainly small pelagic
fish). Moreover, the negative values of the index for the last
part of the time series could be likely related to an impairment of the underlying food web and the ecosystem functioning. The demersal: total catch ratio reflects the increasing
importance of demersal catches mainly due to reduced pelagic
catches (Coll et al., 2005). These results should be viewed in the
context of recent decreasing trends in the NW Mediterranean
pelagic landings and biomasses. After 1994, the official landings have shown a steady decline, mainly due to the decline of
the pelagic fraction, whereas demersal catches seem to have
remained relatively stable since 1983. This would also be in
line with recent evaluations of high risk of ecosystem overfishing sensu Murawski (2000) related with the current fishing
activity in the area (Tudela et al., 2005) and with the ecosystem
effects of fishing reported for the Mediterranean Sea (Tudela,
2004).
This decrease in the pelagic contribution to the catches
(mainly based on organisms with low trophic levels) is also
seen in the analysis of the mean trophic level of the catch
and catches with time, where the TLc increases from 1976 to
2004 and landings decrease. This has also been seen in the
Namibian system (Cury et al., 2005), where a more demersal dominated ecosystem have been observed, resulting from
the decrease and non-recovery of the pelagic component. In
Namibian system the decrease of the FIB index is likely to be
reflecting the collapse of the underlying food webs (Cury et
al., 2005; Heymans, 2004; Heymans et al., 2004; Sumaila et al.,
2004; Willemse and Pauly, 2004). On the contrary, a decrease in
the TLc was found in the Western Mediterranean from 1972 to
1998 (Pinnegar et al., 2003), but when excluding clupeid land-
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e c o l o g i c a l m o d e l l i n g 1 9 8 ( 2 0 0 6 ) 53–70
ing from the analysis a marginally significant increase in the
mean trophic level of capture fishery and aquaculture landings was described.
Similarities between the NW Mediterranean and the
Namibian system are also reflected in the significant consumption by gelatinous zooplankton and the importance of
pelagic-demersal coupling, in remarkable contrast to the other
ecosystems. The importance of gelatinous zooplankton in
both ecosystems appears to be a key difference from the
other ecosystems modelled. The proliferation of jellyfish in
Namibia appeared after the collapse of sardine fisheries in the
1960–1970s and jellies have been dominating the system with
the pelagic goby and the horse mackerel (Heymans et al., 2004;
Roux and Shannon, 2004). The proliferation of jellyfish in the
NW Mediterranean from the 1980s has been also described
(e.g. Buecher, 1999), whilst concern about ecosystem status
is rising due to the decrease of important pelagic resources,
reflected both in biomasses and catches in the system.
The pelagic-benthic coupling is likely to be indicating the
changed state of the Namibian system due to fishing pressure
during the pre-independence period, exacerbated by environmental perturbations and changes (Boyer et al., 2001). This
ecosystem experienced a change from a pelagic-dominated
environment towards a demersal one due to the collapse of
small pelagic fish in the 1960–1970s (Heymans et al., 2004;
Moloney et al., 2005; Roux and Shannon, 2004; Sumaila et al.,
2004; Willemse and Pauly, 2004; Van der Lingen et al., 2006).
The NW Mediterranean could follow suit in the future, whilst
at present the demersal assemblage seems to be growing in
importance relative to the pelagic one (Coll et al., 2005). The
Namibian experience is especially relevant in the context of
the NW Mediterranean because even though Namibia implemented a resource management system to rebuild stocks after
independence in 1990, the small pelagic fish stocks are not
showing clear signs of recovery (Sumaila et al., 2004; Willemse
and Pauly, 2004). Therefore, these similarities between the NW
Mediterranean and Namibian systems should be interpreted
in terms of dynamic trajectories that the Namibian upwelling
has shown due to fishing intensity and the signs that the NW
Mediterranean is showing in that direction.
The present contribution highlights the usefulness of
cross-system comparisons of standardized models. Even if
ecological models are unable to fully capture reality and are
built with values associated with different levels of uncertainty, they are the “best” picture of the ecosystem with
the available information, and the standardization process
helps to minimize the errors associated with the structure
of the model so that the features of the ecosystems can be
revealed and compared. These results support the conclusion
of Moloney et al. (2005) that comparisons of global indices are
useful to generalise ecosystem structure and fishing properties whilst highlighting uncertainties of the model parameters. This is especially important in systems where there is a
deficit of time series data, as in the case of the Mediterranean
Sea.
Moreover, these comparisons are particularly valuable.
Firstly, they serve as a basis for developing and testing hypotheses using dynamic simulations of fishing and
environmental effects along the lines of those examined
in upwelling areas. For example, of special relevance are
the findings that high fishing mortality and environmental anomalies could have driven the Namibian ecosystem
from 1970s till the present and could be related to the collapse of small pelagic fish (Heymans, 2004; Heymans et al.,
2004; Roux and Shannon, 2004). Secondly, a comparative
approach, such as that followed in this study, provides a basis
from which the mechanisms that drive ecosystem changes
and/or regime shifts, and the underlying processes and internal controls operating to sustain ecosystem states, can be
explored.
Acknowledgements
The developers of the ecological models included in the
comparison are acknowledged: Sergio Neira (Chile), Astrid
Jarre (Peru) and Jean-Paul Roux (Namibia). The authors wish
to thank them for agreeing to the use of their published model
results. They also wish to acknowledge all the researchers
involved in the development of the ecological model of the
South Catalan Sea, especially from the Institute of Marine
Science (Barcelona). The Benguela Ecology Programme and
the University of Cape Town are thanked for logistical support and the Spanish Ministry of Education and Science for
financial assistance. An anonymous reviewer is thanked for
suggesting PCA analysis as a means to strengthen our results
and conclusions.
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