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Effect of illegal harvest on apparent survival of Amazon River dolphins (Inia geoffrensis)
Article in Biological Conservation · February 2013 DOI: 10.1016/j.biocon.2012.10.006
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Effect of illegal harvest on apparent survival of Amazon River dolphins (Inia geoffrensis) ⇑ Vanessa J. Mintzer a,b, , Anthony R. Martin c, Vera M.F. da Silva d, Andrew B. Barbour b, Kai Lorenzen b, Thomas K. Frazer a,b a School of Natural Resources and Environment, University of Florida, 103 Black Hall, Box 116455, Gainesville, FL 32611, USA b Fisheries and Aquatic Sciences Program, School of Forest Resources and Conservation, University of Florida, 7922 NW 71st St., Gainesville, FL 32653, USA c Centre for Remote Environments, University of Dundee, 23, Springfield, Dundee DD1 4JE, UK d Instituto Nacional de Pesquisas da Amazônia – INPA/Laboratório de Mamíferos Aquáticos, Av. André Araújo 2936, Manaus, Amazonas 69060-001, Brazil article info abstract
Article history: The Amazon River dolphin (Inia geoffrensis), or boto, is illegally harvested for use as bait in fisheries for the Received 30 June 2012 catfish Calophysus macropterus. To determine the effect of this harvest, we estimated apparent survival for Received in revised form 5 October 2012 a boto population in the central Brazilian Amazon where direct harvest is known to have occurred since Accepted 11 October 2012 2000. For our analysis, we used capture and recapture/resighting data of 528 marked botos over a 17-year Available online 29 November 2012 period (1994–2011). Time-dependent models estimated that apparent survival after the first reports of harvest (/ = 0.899; SE = 0.007) was significantly lower than in years prior to harvest (/ = 0.968; Keywords: SE = 0.009). The decline in apparent survival suggests that current harvest rates exceed conservation lim- Barker model its and may be unsustainable. This issue requires the attention of natural resource managers from all Boto Calophysus macropterus countries of the Amazon basin, as the harvest is widespread and decline in survival could be mirrored Mark-recapture in numerous locales. Bait Ó 2012 Elsevier Ltd. All rights reserved. Poaching
1. Introduction tion, or effects of the harvest (Diniz, 2011). However, in a limited number of cases where such ancillary data has been provided, find- Interactions with fisheries, mostly through incidental capture in ings suggest illegal harvests for bait may have considerable im- fishing gear but also through targeted harvesting are recognized as pacts on targeted populations (e.g. da Silva et al., 2011; Manzur a key threat to aquatic mammal populations (Clapham and Van and Canto, 1997; Sinha, 2002). Waerebeek, 2007; Costello and Baker, 2011; Jefferson and Curry, The Amazon River dolphin (Inia geoffrensis de Blainville, 1817), or 1994; Mangel et al. 2010; Northridge and Hofman, 1999; Perrin boto, has been harvested for use as bait in the fisheries for the catfish et al., 1994; Read, 2008; Read et al., 2006; Robards and Reeves, known as piracatinga, mota, simi, zamurito or mapurite (Calophysus 2011; Vidal, 1993). Perhaps the least well quantified aspect of macropterus Lichtenstein, 1819) (da Silva et al., 2011; Gómez et al., aquatic mammal–fisheries interactions is the harvesting of mam- 2008; Gómez-Salazár et al., 2012; Loch et al., 2009; Portocarrero mals for use as bait, a practice affecting at least 38 species of aqua- Aya et al., 2010a; Shostell and Ruiz-Garcia, 2010; Trujillo et al., tic mammals worldwide (Diniz, 2011). Direct harvest of aquatic 2010b, 2010c). During the last decade, demand for C. macropterus mammals for this purpose is considered an illegal activity in many has grown in Colombia as other species of catfish once popularly nations where it takes place; hence, the prevalence of this illicit consumed in the country have been overfished (Gómez et al., practice and subsequent impact has proved challenging to quan- 2008; Petrere et al., 2004; Trujillo et al., 2010b). In fact, demand is tify. Most studies on the issue only provide information on where such that an international market has emerged with Brazil, Venezu- the illicit harvest occurred without related data on intensity, dura- ela, Peru, and Bolivia, all contributing to the export of this food fish to Colombia, where it is often sold under the name capaz, one of the depleted catfish species from the Magdalena River (Diniz, 2011; ⇑ Corresponding author at: School of Natural Resources and Environment, Gómez et al., 2008; Trujillo et al., 2010b, 2007). More recently, con- University of Florida, 103 Black Hall, Box 116455, Gainesville, FL 32611, USA. Tel.: sumption of C. macropterus has also gained popularity in a few Bra- +1 352 359 5633; fax: +1 352 392 9748. zilian cities where it is marketed as douradinha, a ‘‘new’’ species of E-mail addresses: vjs@ufl.edu (V.J. Mintzer), [email protected] (A.R. Martin), tucuxi@inpa.gov.br (V.M.F. da Silva), snook@ufl.edu (A.B. Barbour), klorenzen@ufl. food fish (da Silva et al., 2011; Trujillo et al., 2010b). Thus, although edu (K. Lorenzen), frazer@ufl.edu (T.K. Frazer). the boto is considered the least endangered of the river dolphins
0006-3207/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.biocon.2012.10.006 V.J. Mintzer et al. / Biological Conservation 158 (2013) 280–286 281
(Inioidea and Platanistidae; Leatherwood and Reeves, 1994; Smith Tefé in Amazonas State, Brazil (Fig. 1). The study area straddles and Smith, 1998), the demand for C. macropterus and related harvest the southern edge of the MSDR. Typical of whitewater floodplain of boto for bait raises substantial concern. or várzea habitat, the MLS has numerous channels and lakes and At present, harvest of botos for bait is known to occur in at least a diverse fish fauna that varies seasonally in response to water twelve locations in four of the five Amazonian countries with C. fluctuations. During May and June, water levels rise 11–15 m and macropterus fisheries (Diniz, 2011), including in and around the the MLS is completely flooded. Lowest water levels typically occur Mamirauá Sustainable Development Reserve (MSDR) upriver of between September and November. These water fluctuations the Brazilian town of Tefé. Although the harvest in this region determine what habitats are available for the boto population. may have started as early as the mid-1990s, the first cases in MSDR While botos may be restricted to the deepest channels and lakes were documented in 2000 (da Silveira and Viana, 2003; Estupiñán during the low water months, they may swim freely and enter sub- et al., 2003). Using C. macropterus landings data from Tefé, da Silva merged forest at high water. Previous estimates suggest approxi- et al. (2011) estimated that roughly 1650 botos are killed in the mately 260 botos occur in or near the 225 km2 MLS year-round area per fishery season. Moreover, the same authors reported a (Martin and da Silva, 2004a, 2004b). Half of these individuals exhi- 10% average annual decline in boto abundance in one of the bit strong site fidelity to the study area, while others are consid- MSDR’s main channels beginning in 2001, a trend coincident with ered transient and visit for shorter periods (Martin and da Silva, the first documented occurrence of the illegal harvest. These find- 2004a). ings highlight the need to further explore the impact of harvesting for bait on the boto population. 2.2. Capture and recapture/resight Herein, we provide estimates of apparent survival for the boto population occurring in and around the Mamirauá Lake System Our capture–recapture and resight data spans 17 years (MLS) at the southern edge of the MSDR. Using mark-recapture/ (January1994–November 2011). These data were collected through resighting models, we explore changes in survival between periods Projeto Boto, a river dolphin research program in the MSDR. of varying harvest pressure. We expected that survival estimates Capture–recapture and marking of botos occurred approximately prior to the documented start of the harvest, i.e. 1994–2000, would 3 weeks each year. With few exceptions, we captured botos with be greater than estimates for the more recent period for which data seine nets during low water at the entrance of the MLS. When cap- were available, i.e. 2000–2011. Additionally, we compared survival tured, we freeze-branded botos with a unique code in areas allow- between the sexes, with the expectation that survival of males ing for maximum contrast and visibility (usually the dorsal fin and would be significantly lower than that of females, based on differ- flank). When previously marked botos were captured, we recorded ences in morphology and behavior. Males are larger (55% heavier their code and rebranded the individual if necessary. Further and 16% longer than females), exhibit high levels of intermale description of capture procedures is available in da Silva and Mar- aggression (Martin and da Silva, 2006), and show preference for tin (2000) and Martin et al. (2006). These data were collected with river habitat outside the protected area boundary (Martin and da the approval of the Instituto Chico Mendes de Conservação da Bio- Silva, 2004b). diversidade (Sistema de Autorização e Informação em Biodiversid- ade #13462-1). 2. Methods In addition to the capture–recapture events, we conducted year-round observational work to provide resightings between 2.1. Study area capture events. We conducted sighting surveys throughout the MLS and surrounding areas, including segments of the main rivers. All data were collected in and around the MLS located at the Due to varying water levels, we did not evenly distribute effort confluence of the Solimões and Japurá rivers, 20 km upriver of across the entire study area. Most areas, however, were surveyed
Fig. 1. Map of study site, the Mamirauá Lake System and surrounding areas, located at the junction of the Japurá and Solimões rivers in the southern segment of the Mamirauá Sustainable Development Reserve in Amazonas State, Brazil (GIS layers: IUCN and UNEP 2010; DCW and GADM downloaded from www.diva-gis.org). 282 V.J. Mintzer et al. / Biological Conservation 158 (2013) 280–286
The Barker model is heavily parameterized; hence, we made a 1700 priori assumptions on how to treat p, r, and R (Table 1). We fixed r as a time-independent (.) parameter because we had few dead 1600 recoveries (only 12 marked carcasses were found during the study period) and therefore lacked sufficient data to inform time-depen- 1500 dent estimates. We defined R and R0 as time-dependent parameters to allow for changes in observation effort through time (Fig. 2). 1400 Both of these parameters were allowed to vary between 3-year intervals (3y) to reduce the parameter count. Capture probability 1300 (p) was treated as a time-dependent parameter (t) because we ex- 1200 pected p to vary according to the environmental conditions of each capture expedition. We created a series of models that allowed /, F, 0 Hours of observational effort 1100 and F (Table 1) to vary with time dependence (t) or independence (.). We also built models that allowed / and F to vary according to 1000 sex (G). Additionally, because the illegal harvest of botos in the MSDR started approximately in 2000, we created models where 2000 2002 2004 2006 2008 / was allowed to vary between two periods: 1994–2000 (i.e. pre- Year harvest) and 2000–2011 (i.e. harvest). The data were partitioned on November 20, 2000, the last day of the 2000 capture expedition. Fig. 2. Hours of observation work conducted in and around the Mamirauá Lake System to provide resight data of Inia geoffrensis between primary capture events. We conducted a median ^c goodness-of-fit test on the global Effort from 1999 to 2009 is displayed showing an increasing trend over time. These model /(G t)p(G t)r(G t)R(G t)R0(G t)F(G t)F0(G t) to as- sighting surveys were typically conducted by 2–3 observers from a 4 m aluminum sess overdispersion (White and Burnham, 1999). We then applied skiff, powered by a 15HP outboard motor. the estimated c (variance inflation factor) to the model set. Values that do not fall within 1 6 c 6 4 indicate a structural lack of fit at least once per week by 2–3 observers from a 4 m skiff. We in- (Burnham and Anderson, 2002). We used Akaike’s Information Cri- cluded additional sightings from other platforms, such as the Pro- terion (AIC) values to rank models (Akaike, 1973). Lower AIC values jeto Boto floating laboratory. When a marked boto was sighted, we indicate models that better explain the variation in the data with recorded its unique code and our percent confidence in the identi- maximum parsimony (Taper et al., 2008). Because we applied the fication (ranging from 50% to 100%, but normally 100%). Since estimated c to our model set, we used the small-sample, ^c-cor- 1999, an average of almost 1500 h of observational work has been rected version of AIC, QAIC , for model ranking. Models with QAIC conducted each year with an increasing trend over time (Fig. 2). c c values that differed by less than 2 were considered equal (Burnham and Anderson, 2002, 2004). We tested for a significant 2.3. Survival model selection and data analysis difference between apparent survival estimates corresponding to different time periods and sex by examining the 95% confidence We estimated apparent survival of the marked population with intervals (CIs) of the /b-values representing the time or sex effect the Barker model (Barker, 1997, 1999) in Program MARK (White (if the 95% CIs of the b-values did not include 0.0, a significant ef- and Burnham, 1999). This model allows for physical capture–re- fect was determined at the a = 0.05 level). capture during discrete primary periods, and continuous resighting data during secondary periods (the intervals between primary periods). We created capture histories for marked individuals cap- 3. Results tured between 1994 and 2011, using resightings made with 100% confidence. We coded for uneven time intervals by scaling We uniquely branded 528 botos (256 females and 272 males). 365 days to equal an interval of length 1. Excluding botos first marked in 2011, we resighted 87.71% of fe- Program MARK estimates seven parameters for the Barker mod- males and 88.93% of males at least once. We recaptured an average el (Table 1)(White and Burnham, 1999). As resightings did not oc- of 25.56 (±17.90 SD) branded individuals each year (Fig. 3). cur over the entire geographic range of the study population; we The median ^c goodness-of-fit test resulted in ^c ¼ 1:162, well estimated apparent survival (/), which includes the effects of emi- within the acceptable range, and we adjusted all model results gration (/ = true survival x (1 - probability of emigration)). We with this value. QAICc strongly supported a model with fully transformed real parameter values [0,1] to b-values [ 1,1] with time-dependent survival estimates (Table 2; Model 1). This model the sin-link or logit-link function for numerical optimization. We estimated a significant difference in F by sex, with estimates of / used simulated annealing for optimization due to a prior simula- = 0.808 (SE = 0.051) for females and / = 0.706 (SE = 0.064) for tion that identified problematic likelihood surfaces leading to con- males (Table 2; Model 1). Capture probability (p) varied consider- vergence failures when using the Newton-Raphson method with ably between years, with an average of p=0.253. This model sug- the Barker model. gested declining apparent survival through time, with an average
Table 1 Barker joint data model parameter definitions.
Parameter Definition
/i The probability that an animal remains alive and available for recapture from i to i +1
pi The probability an animal at risk of capture at i is captured at i
ri The probability an animal that dies in i, i + 1 is found dead
Ri The probability an animal that survives from i to i + 1 is resighted alive some time between i and i +1 0 Ri The probability an animal that dies in i, i + 1 without being found dead is resighted alive in i, i + 1 before it died Fi The probability an animal at risk of capture at i is at risk of capture at i +1 0 Fi The probability an animal not at risk of capture at i is at risk of capture at i + 1 (this definition, as applied in Program MARK White and Burnham (1999), differs from the one provided in Barker (1997) in order to enforce internal constraints) V.J. Mintzer et al. / Biological Conservation 158 (2013) 280–286 283
1.00
250 ) 1 0.95
200 year (
Sex F 150 M
Type 0.90 Count Recapture Apparent survival 100 Resighting
0.85 50 1994 - 2000 2000 - 2011 Time interval
0 Fig. 5. Apparent survival estimates for the pre-harvest (January 1994–November 1995 2000 2005 2010 2000) and harvest period (November 2000–2011) for Inia geoffrensis occurring in and around the Mamirauá Lake System. These estimates were the results of the Year model: /(94-00)(00-11)p(t)r(.)R(3y)R0(3y)(F(G)F0(t). The difference in apparent survival probability between periods was found to be significant at the a = 0.05 Fig. 3. Capture, recapture, and resight counts from 1994 to 2011 of Inia geoffrensis level. occurring in and around the Mamirauá Lake System. Capture counts (solid lines) displayed by sex represent the cumulative number of individuals marked. Approx- imately 3 weeks each year were dedicated to the capture–recapture and marking of botos. Observational work was conducted between primary capture events to obtain resighting data. and harvest periods, we found a significant difference (pre-harvest: / = 0.968, SE = 0.009; harvest period: / = 0.899, SE = 0.007) (Ta- ble 2; Model 5; Fig. 5). Table 2 The model including sex as a group attribute of survival (Ta- QAICc table from Barker survival model results. ble 2; Model 8) estimated a non-significant difference in annual Model QAIC QAIC QAIC weight k c c c apparent survival between females (/ = 0.919, SE = .008) and males 1. /(t)F(G)F0(t) 6097.76 0.00 0.71 69 (/ = 0.904, SE = 0.009). Moreover, no significant difference was 0 2. /(t)F(.)F (t) 6099.87 2.11 0.25 68 found between apparent survival of females and males for the 3. /(t)F(t)F0(t) 6105.55 7.79 0.01 84 4. /(t)F(t)F0(.) 6105.57 7.80 0.01 69 pre-harvest (females: / = 0.973, SE = 0.011; males: / = 0.963, 5. /(94-00)(00-11)F(G)F0(t) 6106.82 9.06 0.01 52 SE = 0.014) or harvest period (females: / = 0.906, SE = 0.009; 6. /(G (94-00)(00-11)F(G)F0(t) 6109.41 11.65 0.00 54 males: / = 0.891, SE = 0.010) (Table 2; Model 6). 7. /(.)F(G)F0(t) 6128.11 30.36 0.00 51 8. /(G)F(G)F0(t) 6128.53 30.78 0.00 52 4. Discussion The parameter of primary interest was apparent survival (/). Time-dependence (t) and sex effect (G) were represented with the associated symbols. Parameters p, r, R, and R0 were fixed through a priori assumptions (see Section 2.3). Number of esti- 4.1. Model performance mated parameters (k) is listed for each model. The Barker model fit the data well ð^c ¼ 1:162Þ and accounted for variable sighting effort with time-dependent estimates of the res- annual apparent survival of 0.912 over the study’s duration (Fig. 4). ighting parameter, R. Although botos are known to occasionally When estimating apparent survival separately for the pre-harvest emigrate from the study area for several years, it is unlikely this temporary emigration induced a major negative bias in apparent survival estimates as the Barker model is robust to emigration 1.0 due to its parameterizations (Barker, 1997; Horton and Letcher, )
1 2008). Additionally, the ^c value near 1 suggested a lack of major r
a 0.9 model assumption violations. e y ( 0.8
val 4.2. Survival estimates 0.7 Survi
t Our results indicate that annual apparent survival significantly
n 0.6 decreased by 0.069 after the documented start of the harvest.
pare 0.5 The apparent survival estimate of 0.968 for the pre-harvest period
Ap is consistent with boto life history, which is characterized by traits 0.4 associated with high annual survival such as slow maturation and 1995 2000 2005 2010 birth intervals of approximately 3 years (see da Silva, 2008). This Year apparent survival estimate is comparable to survival estimates of non-harvested populations of Tursiops truncatus (Speakman et al., Fig. 4. Apparent survival estimates for 1994–2011 for Inia geoffrensis occurring in 2010), Tursiops aduncus (Mansur et al., 2012), and Orcinus orca and around the Mamirauá Lake System. Apparent survival estimates displayed were the results of the Barker model where / was treated as a fully time-dependent (Matkin et al., 2012) which range from 0.951 to 0.990. However, parameter: /(t)p(t)r(.)R(3y)R0(3y)(F(G)F0(t). annual apparent survival estimates for the full study duration (/ 284 V.J. Mintzer et al. / Biological Conservation 158 (2013) 280–286
= 0.912) and for the harvest period (/ = 0.899; SE = 0.007) are low- and Best, 1996). Many of the boto carcasses recovered by Projeto er than most published survival estimates of other odontocetes, Boto during the study period showed evidence of entanglement suggesting a negative effect of harvest on the population. (da Silva and Martin, 2010; Martin et al., 2004). Furthermore, We did not expect equal apparent survival between the sexes. observations suggest fishers kill botos intentionally because they The species is sexually dimorphic, and males exhibit corresponding regard them as competitors, and because botos cause damage to high levels of intermale aggression that may lead to life threaten- fishing nets (da Silva and Best, 1996; da Silva and Martin, 2010; ing injuries (Martin and da Silva, 2006). Furthermore, when a wide Loch et al., 2009). The number of artisanal and commercial fleets range of habitats is available during high water, females show pref- in the study region has remained relatively stable in the last two erence for várzea habitats within the MLS, while males are found decades after experiencing a period of fast growth in the 1970s more frequently in the main rivers outside of the protected area and 1980s (Almeida et al., 2003; Barthem, 1995; Best and da Silva, boundary (Martin and da Silva, 2004b). Thus, in theory, males are 1989). Thus, although fishery interactions other than boto harvest at a higher risk of being harvested because they spend more time for bait are likely causes of additive mortality for our study popu- outside protected waters. For these two reasons, we expected low- lation (Martin et al., 2004), there is no indication that such interac- er apparent survival in males. It is possible the harvest is exerting tions have increased considerably since the onset of this study, and selective pressure on females (if, for example, females are easier to are therefore not likely to explain the decrease in apparent sur- catch), and increasing mortality of females to the extent that sur- vival. However, this threat should be carefully considered in con- vival of females and males have equalized. However, our results servation and management plans, as it was the primary cause of did not show a significant difference between male and female the functional extinction of the Yangtze River dolphin (Turvey apparent survival during the pre-harvest period, suggesting that et al., 2007) and changes in the size, gear, or techniques of Amazo- similar survival probability between the sexes is a natural nian fishing fleets could lead to considerable increases in boto characteristic. mortality. The significant decline in apparent survival estimated in this 4.3. Significance of the harvest and conservation implications study suggests a need for conservation measures aimed at decreas- ing the boto harvest. Currently, two main tools are in place that Assuming declines in apparent survival were due to lethal as should aid in the protection of the study population. First, the boto opposed to sub-lethal effects (e.g. changes in behavior), the results is protected by Brazilian federal laws, predominantly Law 7.643 indicate that mortality of the study population more than doubled (1987) that makes it illegal to kill or harass cetaceans in jurisdic- between the two time periods. Dividing the harvest period appar- tional waters. However, enforcement of natural resource protec- ent survival estimate by the pre-harvest period estimate, we calcu- tion laws in the Amazon is challenging, and efforts are often late a reduction in survival corresponding to an annual harvest rate compromised as a consequence of budget constraints, understaff- (h) of 0.071. The Potential Biological Removal (PBR) method, com- ing, and other institutional deficiencies (McGuire and Aliaga- monly applied to cetaceans, calculates the level of mortality that Rossel, 2010; Peres and Terborgh, 1995; Trujillo et al., 2010c; will inhibit a stock from reaching or maintaining its optimum sus- Utreras et al., 2010). This appears to be the case in our study region, tainable population (Wade, 1998). According to the PBR approach, where only four federal officials have been available to enforce the maximum allowable harvest rate (hmax) is equal to 1/2Rmax (see conservation related regulations in an area that exceeds 2 Dillingham and Fletcher, 2008), where Rmax is defined as the max- 251,000 km (Peres and Lake, 2003). Furthermore, lack of aware- imum annual recruitment rate (Wade, 1998). If we apply an Rmax of ness by fishers of existing legislation has been noted as an imped- 0.04, the default value used for cetaceans (Wade, 1998), the hmax iment to the conservation of botos (e.g. McGuire and Aliaga-Rossel, would equal 0.02, a value about three times smaller than our esti- 2010; da Silva and Martin, 2010). mated h. If we apply a higher Rmax value of 0.06 (theoretically pos- The second major conservation mechanism is the partial spatial sible in odontocetes with high survival, Reilly and Barlow, 1986; protection provided by the MSDR. Although MSDR was not created
Wade, 1998), the estimated h is still more than double the hmax with the specific purpose of protecting river dolphins, the reserve of 0.03 calculated by the PBR rule. This comparison suggests the includes important river dolphin habitat. To our knowledge, the current harvest rates exceed conservation limits commonly ap- harvest does not occur in the section of our study area that falls plied to cetaceans and could lead to depletion of the population within the MSDR (see Fig. 1). Our results indicate that both females (Dillingham and Fletcher, 2008). (F = 0.808; SE = 0.051) and males (F = 0.706; SE = 0.064) exhibit It is possible that factors other than illegal harvest could ex- high fidelity to our study area; therefore, the protected area should plain, in part, declines in boto survival. Throughout its geographic function by protecting the population at least part of the time (pri- range, dam construction, pollution, entanglement in fishing gear, marily during rising water when botos enter the várzea, Martin and habitat degradation, boat traffic, and depleted prey resources have da Silva, 2004b). If the MSDR did not exist, it is likely the mortality been identified as important anthropogenic threats to the species of our study population would have been higher in the harvest per- (Best and da Silva, 1993; 1989; da Silva and Best, 1996; Gómez- iod. Nevertheless, our results imply that the spatial protection in Salazár et al., 2012; Leatherwood and Reeves, 1994; Martin et al., our study area, with its current design and state, may not be suffi- 2004; McGuire and Aliaga-Rossel, 2010; Portocarrero Aya et al., cient in and of itself to protect this population. 2010a; Reeves and Leatherwood, 1994; Rosas and Lehti, 1996; Marine protected areas have been shown recently to be effective Smith and Smith, 1998; Trujillo et al., 2010b, 2010c; Utreras in increasing survival rates of small cetaceans by reducing fishery– et al., 2010; Vidal, 1993). However, due to the location of the study, dolphin interactions (Gormley et al., 2012). These findings suggest we suggest that aside from illegal harvest, only entanglement in that properly designed protected areas in the Amazon basin might fishing gear is likely to substantially affect mortality of the study translate into similar benefits for river dolphins. As proposed by the population. South American River Dolphin Protected Area Network (SARDPAN), Although incidental mortality in fishing gear has not been stud- protected areas created or improved explicitly for river dolphin ied in most areas in the Amazon, it is known that seine nets and conservation could have the potential to reduce the impact of fish- gillnets pose a significant threat to botos (e.g. da Silva and Best, ery–river dolphin interactions, as well as help conserve vulnerable 1996; Leatherwood and Reeves, 1994). In the central Amazon, freshwater habitats largely underrepresented in South American the lampara seine is the most lethal type of net for botos and ac- protected areas (Portocarrero Aya et al., 2010b). Although much re- counts for over 80% of deaths caused by entanglement (da Silva mains to be understood about river dolphin ecology, several studies V.J. Mintzer et al. / Biological Conservation 158 (2013) 280–286 285 have provided valuable information on boto habitat preference and Acknowledgements movement that could prove useful in designing new protected areas or improving existing ones (e.g. Gómez-Salazár et al., 2012; This study was part of Projeto Boto, a cooperative agreement Leatherwood et al., 2000; Martin and da Silva, 2004a, 2004b; Martin between the National Amazon Research Institute-INPA/MCTI and et al., 2004; McGuire and Henningsen, 2007; McGuire and the Mamirauá Sustainable Development Institute – MSDI-OS/ Winemiller, 1998). In our study site, for example, extending spatial MCT. It was completed with support from the School of Natural Re- protection to adjacent areas where botos aggregate (e.g. mouths of sources and Environment, the Fisheries and Aquatic Sciences Pro- channels that connect to the main rivers), could benefit the study gram, and the Tropical Conservation and Development Program population. The effectiveness of new or enhanced conservation at the University of Florida; the Society for Marine Mammalogy; measures can be monitored and evaluated using the baseline sur- and the Associação Amigos do Peixe-Boi. We thank the many peo- vival estimates provided in this study. ple who contributed to the fieldwork on which this study is based, Although improvements made to enforcement and spatial pro- including the numerous Projeto Boto interns that collected the data tection could prove to be beneficial, numerous challenges will have analyzed. We also extend a special thanks to Dr. Claudia Peñaloza to be overcome to decrease boto harvest. With administrative and Dr. William Pine for their valuable modeling advice. ABB was boundaries being crossed, the remoteness of the harvest sites, supported by a National Science Foundation Graduate Research and high economic incentives, both local and international action Fellowship under Grant No. DGE-0802270. The views expressed and cooperation will be required. Broad scale initiatives like the are those of the authors and do not necessarily reflect the views SARDPAN could facilitate the necessary international collaboration of the above mentioned organizations. and the implementation of new and improved conservation mech- anisms. Moreover, it is imperative that all drivers of the harvest be References considered when formulating solutions, including the troubling status of the preferred food catfish species in Colombia, as well Akaike, H., 1973. Information theory and an extension of the maximum likelihood principle. In: Petrov, B.N., Caski, F. (Eds.), Second International Symposium on as the socio-economic conditions of fishers at the harvest sites. Information Theory. Akademiai Kiado, Budapest, pp. 267–281. With this in mind, solutions of varying scope and scale should be Almeida, O.T., Lorenzen, K., McGrath, D.G., 2003. Commercial fishing in the Brazilian simultaneously explored. 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macropterus) na Reserva de Desenvolvimento Sustentável Mamirauá. Relatório (Brachyplatystoma filamentosum Lichtenstein). Rev. Fish Biol. Fisher. 14, 403– Técnico. Ministério da Ciência e Tecnologia, Instituto de Desenvolvimento 414. Sustentável Mamirauá, Tefé, Brasil.
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