Efectos de pesca sobre elasmobranquios en el Pacifico colombiano.
Efeitos sobre a pesca nos elasmobranquios no Pacifico colombiano
Over die last 50 years, sustained fishing activity worldwide and die steady degradation of habitats have had a wide range of impacts on ecosystems; reflecting in changes of abundance, spatial distribution, productivity, and structure of those communities under exploitation (Hall 1999, Jackson et al. 2001, Myers & Worm 2005, Lotze et al. 2006, Navia et al. 2012). This impact on die structure and function of communities has been widely documented, quantified, and found to impair the maintenance of goods and services that ecosystems provide to humans (Worm et al. 2006, Branch et al. 2010, Lotze et al. 2011).
There are ecological limits (especially related to the energetic balance within food webs) that explain die low number of top predators that may be extracted from a marine environment (Perez-Espana et al. 2006). Regardless of known implications, most fisheries tend to focus on these top predators (Branch et al. 2010). Pauly et al. (1998) documented the results of such extractive activities as "fishing down marine food webs": a gradual change in landings from long-lived, upper trophic level species to short-lived, lower trophic species. This phenomenon was observed at both a global (Pauly et al. 1998) and local scale in Thailand (Christensen 1998), Canada (Pauly et al. 2001), the Mediterranean Sea (Pinnegar et al. 2003), Chile (Arancibia & Neira 2005), Uruguay and Argentina (Jaureguizar & Milessi, 2008), India (Bhathal & Pauly 2008), and Brazil (Freire & Pauly 2010). Conversely, some authors suggest that the decline in mean trophic levels of fisheries landings is due to increased exploitation of the lower trophic levels of marine food webs (Essington et al. 2006). This effect, called "fishing through the food web", has been recorded in North Pacific fisheries (Litzow & Urban 2009).
Sharks and rays are associated with many fisheries worldwide and more than half of die annual catch is through by-catch (Bonfil 1994, Stevens et al. 2000). Consequently, these species are listed as highly vulnerable to exploitation (Holden 1974, Walker 1998) and considered to be among the most endangered vertebrate groups in the world (Stevens et al. 2000, Dulvy & Reynolds 2002, Dulvy et al. 2014). Numerous studies have recorded high reduction rates of elasmobranchs over the past 50 years in many fisheries (e.g. Graham et al. 2001, Coelho et al. 2003, Baum et al. 2003, Myers & Worm 2005). Moreover, it has been proposed that the reduction or disappearance of sharks considered as top predators may have strong effects on die function and structure of the marine food webs (Stevens et al. 2000, Myers et al. 2005, Navia et al. 2010, Navia 2013).
The central and southern Pacific coast of Colombia have been key areas to the domestic shrimp industry since 1960 (de la Pava & Mosquera 2001). Several species of sharks and rays are caught as by-catch and some are commercially exploited (Mejia-Falla & Navia 2011). Despite the history of the shrimp industry in this area, official records of catch efforts and landings are poor (Rueda et al. 2006, Mejia-Falla & Navia 2011). As a result, data required for the design of an appropriate fishery management plan is lacking. Furthermore, trawling impacts and changes to the fish assemblages over time in this area have not been assessed. Based on industrial shrimp trawling fleet data, the goal of our research is to evaluate the changes in the elasmobranch assemblage structure between 1995 and 2007 and determine possible effects of trawling on the assemblage.
Materials and methods
Monitoring of onboard shrimp vessels trawling between Golfo de Tortugas and Ensenada de Guapi (Colombian Pacific coast) was conducted every October in 1995, 2001, 2003, 2004 and 2007 (Figure 1). To standardize efforts, all fishing operations were carried out by the same vessel (M/N Arraijan) using die same fishing gear (trawl nets; each 27 m in length, mesh size 2 1/4"); between 11.6 and 31.0 m deep (Table 1), in die same area (Figure 1), and under procedures traditionally used by commercial fishing vessels: i.e. trawling three hours during die day and six hours at night. Following each haul, sharks and rays were identified and recorded to species level; sex, size (total length, Lt in cm), weight (g), and number of specimens per species.
Relative abundance (catch per unit effort) was calculated by dividing the total number of specimens of each species by the total number of hours per haul in die sampling year. Temporal variations in these relative abundances were analyzed using ANOVA (Kruskal-Wallis test, Zar 1999) and differences in species richness over time were evaluated using Cochran's Q test (Zar 1999). Differences in species catch size (Lt) between sampling years were evaluated with a nonparametric ANOVA (Zar 1999).
[FIGURE 1 OMITTED]
To analyze changes in die elasmobranch assemblage attributed over time, we calculated die indices of heterogeneity, diversity, and equitability (Krebs 1999); using die software Ecological Methodology (Krebs 2001). The nonparametric Simpson's index (D) was used to compare heterogeneity, and respectively, die complement index (1-D) towards diversity ranging from 0 (low diversity) to 1 (maximum diversity). Shannon diversity index (H') was also used to measure diversity. Equitability was quantified using Simpson's measure ([E.sub.1/D]), which ranges from 0 to 1 and is not affected by rare species (Krebs 1999). These indices accurately describe the structural and functional attributes of fish communities (Piet & Jennings 2005) and there is available theoretical and empirical evidence that they can measure the impact of fishing (Rochet & Trenkel 2003).
Significant differences ([alpha] = 0.05) between the attributes of the assemblage were evaluated by applying die Hutchenson's t test (1970) to the Shannon diversity index values, under die null hypothesis that the diversity of the samples were not different. This test was calculated as t = ([H'.sub.1] - [H'.sub.2])x[(Var [H'.sub.1] + Var [H'.sub.2]).sup.-0.5], where [H'.sub.1] and [H'.sub.2] are the Shannon diversity indices of samples 1 and 2, respectively, and Var [H'.sub.1] and Var [H'.sub.2] are the variances of each of the indices (Magurran 1988).
To evaluate die effects of trawling on die trophic structure of die ecosystem in question, the historical changes of die mean trophic level (T[L.sub.m]) of the elasmobranch assemblage were analyzed. To quantify die trophic level of die species, studies on feeding habits carried out in die same studied area were used (Gomez et al. 2003, Mejia-Falla et al. 2006, Navia et al. 2006, 2007, 2011, Lopez-Garcia et al. 2012, Navia 2013). For those species whose diet has not been studied in this area, we used studies conducted in die Eastern Tropical Pacific (ETP) (Valadez-Gonzalez 2007). Literature on die diet of Carcharhinus lencas and C. porosas in die ETP was not available. Thus, we used trophic level values calculated by Cortes (1999): TL = 4.3 for C. lencas and TL = 4.1 for C. porosas. Trophic levels (T[L.sub.k]) were quantified by applying die formula proposed by Cortes (1999) and reviewed by Ebert & Bizarro (2007).
T[L.sub.k] = 1 + [[n.summation over (j=1)] [P.sub.j] x T[L.sub.j]] (1)
where [P.sub.j] is die fraction of prey i in the diet of predatory; T[L.sub.j] is die trophic level of prey i in the diet of predatory. We used die reference values of trophic levels for prey suggested by Cortes (1999), Ebert & Bizarro (2007) and Lopez-Garcia et al. (2012).
The mean trophic level (T[L.sub.m]) of the assemblage for each sampling year was calculated according to Pauly & Palomares (2005):
T[L.sub.m] = [summation] T[L.sub.k] [Y.sub.ij]/[summation] [Y.sub.ij] (2)
Where T[L.sub.k] is the mean trophic level of the species or group k, and [Y.sub.ij] is die catch of species or group i in year y
A total of 90 net hauls and 396 hours of fishing were monitored; a total of 1711 specimens of 19 species (7 sharks and 12 rays) were captured (Table 2). The number of species varied between 7 (1995) and 12 (2003, 2007). No significant differences were found in occurrence (Cochran, Q = 4.30, df = 4, p = 0.367) but there were significant differences in relative abundance (CPUE) between studied years (Kruskal-Wallis test, H = 3.01, df = 4, p = 0.045). For example, in sharks, the relative abundance of C. porosas changed from 0.24 in 1995 to 0.01 in 2001, and Mustelus lunulatus changed from 0.48 in 2001 to 0.01 in 2007. Average catch size decreased during diese years from 67.6 to 55.0 cm Lt. Regarding batoids, Dasyatis longa changed from a relative abundance of 0.27 and average catch size of 150.6 cm Lt in 2001 to 0.09 and 129.7 cm Lt, respectively, in 2007. On die other hand, the index of relative abundance of some small sized rays with no commercial value, such as Narcine entemedor, N. leoparda and Urottygon aspidura increased; while the catch sizes declined (Table 2).
There was no clear trend in equitability and heterogeneity over time. Between 1995 and 2001, diese attributes increased, whereas between 2001 and 2007, they decreased. Either way, this suggests an alteration of die species richness and abundance during die research period. Species richness of elasmobranchs differed significantiy between all sampled years (Table 3), indicating considerable variation in the assemblage composition during the studied period.
Over the years studied, die number of species increased (7 to 12) while the T[L.sub.m] decreased slightiy (Figure 2). As an exception to this trend, 11 species were collected in 2001 and the T[L.sub.m] was the second highest of all sampled years. This was due to die catch of Zapteryx xyster and Raja velezi, resulting in a TL higher than the other rays in die study, and only occurred in that year. The T[L.sub.m] of die assemblage ranged from 3.61 in 1995 to 3.55 in 2007. Species from the top predator functional level (TL [greater than or equal to] 4.0), such as those from the Sphyrnidae and Carcharhinidae families, were not recorded during any of the fishing trips in 2007.
[FIGURE 2 OMITTED]
Changes in the elasmobranch assemblage detected during this study included alteration of assemblage attributes (heterogeneity, diversity, and equitability), modification of die relative abundance of most species, a slight reduction in mean trophic level, reduction in average catch size, and a substantial reduction in upper trophic level predators. For example, M. lunulatus and D. longa, considered the most important commercial elasmobranch species in die late 1990s on die Pacific coast (Zapata et al. 1999), showed a significant downward trend in abundance and average catch size. According to Piet & Jennings (2005), this effect on average catch size and species composition is directly linked to changes in die trophic structure of the community.
Thirty of the 48 largest marine ecosystems on die planet have shown reductions in trophic levels (Essington et al. 2006). Although tire results of this study do not demonstrate tire removal of top predators from die studied area, they do show that species with TL [greater than or equal to] 4 decreased significantly. In contrast, tire intermediate trophic level species increased in abundance over die same period. These results are consistent with fisheries statistics for die Colombian Pacific coast (Mejia-Falla & Navia, 2011), which showed a reduction in average catch sizes and an increase in abundance of juvenile and intermediate TL species. This suggests a significant reduction or even possible loss of the ecological function of top predators, especially those at the adult stage, which could impact, in the short or medium term, the structure of the food web in the study area. This is especially important considering top predators were identified as the most important species in keeping the balance of an ecosystem's structure (Navia et al. 2010, 2012, Navia 2013).
The rate of decline of trophic levels in this assemblage (0.0035 [year.sup.-1]) is lower than those reported in other studies (e.g. Pauly et al. 1998, Pinnegar et al. 2003, Milessi et al. 2005; Jaureguizar & Milessi 2008). However, it is worth noting drat in just 5 years of sampling, the T[L.sub.m] of the assemblage showed a reduction. This might reflect the significant impact drat trawling has had on the species within the studied area even though this fishery does not catch high trophic level shark species on a regular basis.
The increase of species richness and evenness between 1995 and 2007 suggests tire occurrence of additional elasmobranchs species that were not associated with tire shrimp trawl fishery in earlier years. This change in tire assemblage composition is tire result of increased catch frequency of intermediate or lower trophic level species including N. leoparda, N. entemedor, Rhinobatos lencotiynchns, and U. aspidura (TL < 3.7).
According to Piet & Jennings (2005), the assumption that assemblages have had an integrated response to fishing pressures implies that all fish community indicators should decrease. This has been observed in shallow water fisheries of the North Sea (Piet & Jennings 2005) and coastal shrimp fisheries along tire Pacific coast of Colombia in tire latter years of this study. It has been argued that community indices and trophic levels calculated from landings in some countries are biased by economic interests towards targeted species (Bianchi et al. 2000, Essington et al. 2006). Consequentiy, changes in community indices and trophic levels are not reliable indicators of tire impact on fisheries. This work, however, was based on direct monitoring onboard fishing vessels and assessed all tire elasmobranch species captured (regardless of their economic value). Therefore, this study provides accurate and reliable information on tire performance of both indicators and assemblage attributes.
The changes in the assemblage found in this study; i.e. reduction in mean trophic level, increased richness of small species, and changes in their abundance; suggests fishing effects some of tire structural and functional parameters of the trophic networks: proportion of top predators, links density, and predatory function. If this trend continues and extends to other functional groups, tire upper trophic levels of tire food web within tire study area will experience overexploitation. Consequentiy, tire network structure will be simplified and the productivity and biomass will decrease, as has been recorded within many other ecosystems highly impacted by fishing (Coll et al. 2008, 2009a, Barausse et al. 2009, Morissette et al. 2009, Navia et al. 2012). Essington et al. (2006) suggests tirat tire reduction observed in trophic levels of an ecosystem can be influenced by tire depletion of upper trophic levels (e.g. fish) and tire creation of alternative fisheries of lower trophic levels (e.g. invertebrates). However, this is not tire case in Colombian Pacific fisheries because despite tire reduction in yields from traditional fisheries, no new alternative fisheries have been developed.
In tire Pacific Region, the shrimp fishery began in the 1960s and peaked in 1974 when there were 138 boats in this fishery (de la Pava & Mosquera, 2001). In tire mid-1980s, tire shrimp crisis began, which led to tire continuous reduction of vessels until 2007 when there were only 29 active (Navia et al. 2008). Despite this drastic reduction in fishing activity, the proportion of by-catch (including elasmobranchs) versus shrimp in the Colombian Pacific Ocean between 2004 and 2005 was around 14:1 kg (Rueda et al. 2006), which was well above the overall average of 10:1 kg in other world tropical fisheries (Hall et al. 2000, Lewinson et al. 2004).
Our results suggest that trawling has possibly impacted die structure and composition of die elasmobranch assemblage in die Colombian Pacific coastal waters. This has led to die reduction of individuals in upper trophic levels, potentially contributing to instability of the food web and/or alteration of ecosystem control mechanisms. Impacts on die elasmobranch assemblage may also trigger gradual changes that could lead to a new organization of die food web as has been suggested for various marine ecosystems (Chen et al. 2008, Coll et al. 2009a, 2009b, Duan et al. 2009, Andersen & Pedersen 2010, Scheffer 2010, Lotze et al. 2011). These changes to die Colombian Pacific ecosystems could make this food web particularly sensitive to die loss of top predators, as suggested by Navia et al. (2012).
The authors wish to thank A. Martinez and his crew for facilitating onboard monitoring from die vessel "Arraijan". They also diank A. Tobon, LF. Payan and S. Hleap for dieir assistance conducting field work. They diank Y Ramirez-Luna and P. Sargent for translating the manuscript into English. This project was supported by Conservacion Internacional and WWF Colombia. Idea Wild donated equipment for field and laboratory work.
Conflicts of interest
The authors declare diat diere are no conflicts of interest. References
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Andres Felipe Navia (1) *, Paola Andrea Mejia-Falla (1)
Juan Carlos Salcedo-Reyes
(1.) Fundacion colombiana para la investigacion y conservacion de tiburones y rayas SQUALUS. Carrera 60A No.11-39, Cali, Colombia
Published on line: 20-01-2016
Electronic supplementary material: N/A
Andres Felipe Navia
His main research interest is directed to the evaluation of the ecological function of elasmobranchs in marine ecosystems, and the determination of the relationship between life history characteristics of these species and their vulnerability, with emphasis on reproduction and age. He recently initiated studies on the essential habitats of elasmobranchs and ecological processes that determine their richness and distribution.
Paola Andrea Mejia-Falla
Her research focused on the life history strategies and demographics of elasmobranchs, as well as to assess the effect of fishing on such population parameters. She has conducted research on reproduction, age and growth of various species of elasmobranchs, both marine and freshwater. She has initiated studies on essential habitats of elasmobranches and on the macro-ecological processes that determine the distribution of these species.
Table 1. Number of trips and trawling depths per year carried out onboard the vessel M/N Arraijan. Year Number Depth Minimum Maximun Average 1995 18 5.5 35.0 18.0 2001 24 8.5 59.4 31.3 2003 16 4.9 25.2 11.6 2004 18 3.6 52.2 20.0 2007 19 10.8 27.0 19.2 Table 2. Trophic level (TL), catch per unit effort (CPUE, used as relative abundance synonymously) and average catch sizes ([L.sup.T] in cm) of elasmobranchs caught during sampled years. Number of effective sampling hours per year is indicated in brackets. 1995 (76) 2001 (87) TL n CPUE LT n CPUE LT SHARKS Family Triakidae Mustelus lunulatus 3.8 3 0.04 36.0 42 0.48 67.6 Family Carcharhinidae Carcharhinus porosus 4.1 18 0.24 41.0 1 0.01 40.0 Family Sphyrnidae Sphyrna corona 4.1 0 -- -- 0 -- -- Sphyrna lewini 4.4 3 0.04 42.5 6 0.07 53.0 Sphyrna media 3.9 0 -- -- 3 0.03 58.9 Sphyrna tiburo 3.2 2 0.03 37.0 3 0.03 83.2 RAYS AND SKATES Family Narcinidae Narcine entemedor 3.7 0 -- -- 12 0.14 58.3 Narcine leoparda 3.6 0 -- -- 0 -- -- Family Rhinobatidae Rhinobatos leucorhynchus 3.5 16 0.21 32.5 28 0.32 55.4 Zapteryx xyster 3.5 0 -- -- 54 0.62 55.4 Family Rajidae Raja velezi 3.6 0 -- -- 14 0.16 69.0 Family Urotrygonidae Urotrygon aspidura 3.7 61 0.79 34.5 38 0.44 32.9 Urotrygon rogersi 3.5 40 0.54 22.9 0 -- -- Urotrygon spp. 3.6 0 -- -- 0 -- -- Family Dasyatidae Dasyatis longa 3.8 0 -- -- 24 0.27 150.6 Dasyatis spp. 3.8 0 -- -- 0 -- -- Family Gymnuridae Gymnura marmorata 3.7 0 -- -- 0 -- -- Family Myliobatidae Aetobatus narinari 3.9 0 -- -- 0 -- -- Species richness 7 11 Total number captured 143 225 2003 (63) 2004 (60) n CPUE LT n CPUE LT SHARKS Family Triakidae Mustelus lunulatus 3 0.05 53.3 0 -- -- Family Carcharhinidae Carcharhinus porosus 11 0.17 37.6 1 0.02 38.0 Family Sphyrnidae Sphyrna corona 7 0.11 52.8 0 -- -- Sphyrna lewini 0 -- -- 1 0.02 88.0 Sphyrna media 0 -- -- 0 -- -- Sphyrna tiburo 4 0.06 40.5 3 0.05 97.5 RAYS AND SKATES Family Narcinidae Narcine entemedor 1 0.02 60.0 2 0.03 37.4 Narcine leoparda 4 0.06 25.7 3 0.05 -- Family Rhinobatidae Rhinobatos leucorhynchus 10 0.16 49.1 16 0.27 59.0 Zapteryx xyster 0 -- -- 0 -- -- Family Rajidae Raja velezi 0 -- -- 0 -- -- Family Urotrygonidae Urotrygon aspidura 66 1.04 30.9 13 0.22 23.9 Urotrygon rogersi 0 -- -- 0 -- -- Urotrygon spp. 0 -- -- 3 0.05 24.1 Family Dasyatidae Dasyatis longa 3 0.05 127.7 10 0.17 125.3 Dasyatis spp. 1 0.02 149.0 -- -- Family Gymnuridae Gymnura marmorata 4 0.06 39.4 2 0.03 26.9 Family Myliobatidae Aetobatus narinari 1 0.02 185.3 0 -- -- Species richness 12 10 Total number captured 115 54 2007 (110) n CPUE LT SHARKS Family Triakidae Mustelus lunulatus 1 0.01 55.0 Family Carcharhinidae Carcharhinus porosus 0 -- -- Family Sphyrnidae Sphyrna corona 0 -- -- Sphyrna lewini 0 -- -- Sphyrna media 0 -- -- Sphyrna tiburo 2 0.02 -- RAYS AND SKATES Family Narcinidae Narcine entemedor 95 0.86 37.8 Narcine leoparda 196 1.78 21.9 Family Rhinobatidae Rhinobatos leucorhynchus 67 0.61 38.9 Zapteryx xyster 0 -- -- Family Rajidae Raja velezi 1 0.01 -- Family Urotrygonidae Urotrygon aspidura 145 1.32 28.5 Urotrygon rogersi 47 0.43 24.8 Urotrygon spp. 607 5.52 42.5 Family Dasyatidae Dasyatis longa 10 0.09 129.7 Dasyatis spp. -- -- Family Gymnuridae Gymnura marmorata 2 0.02 55.5 Family Myliobatidae Aetobatus narinari 1 0.01 124.5 Species richness 12 Total number captured 1174 Table 3. Comparison of Shannon diversity index values ([H.sub.'1] and [H.sub.'2]) between sampling years and statistical significance ([alpha] = 0.05) using the Hutchenson's t test, d.f = degree freedom, p = probability. Years [H'.sub.1] [H'.sub.2] d.f [t.sub.table] 1995 vs 2001 1.17 2.02 210 1.97 1995 vs 2003 1.17 1.59 241 1.97 1995 vs 2004 1.17 1.89 142 1.97 1995 vs 2007 1.17 1.72 168 1.97 2001 vs 2003 2.02 1.59 149 1.98 2001 vs 2004 2.02 1.89 73 2.00 2001 vs 2007 2.02 1.72 381 1.96 2003 vs 2004 1.59 1.89 154 1.96 2003 vs 2007 1.59 1.72 126 1.96 2004 vs 2007 1.89 1.72 60 2.00 Years [t.sub.cal] P 1995 vs 2001 17.80 < 0.05 1995 vs 2003 6.21 < 0.05 1995 vs 2004 11.05 < 0.05 1995 vs 2007 12.33 < 0.05 2001 vs 2003 7.75 < 0.05 2001 vs 2004 2.48 < 0.05 2001 vs 2007 13.12 < 0.05 2003 vs 2004 4.21 < 0.05 2003 vs 2007 2.49 < 0.05 2004 vs 2007 3.35 < 0.05
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|Title Annotation:||ORIGINAL ARTICLE|
|Author:||Felipe Navia, Andres; Mejia-Falla, Paola Andrea|
|Publication:||Revista Universitas Scientarum|
|Date:||Jan 1, 2016|
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