- Top of page
- Materials and methods
- Supporting Information
One of the most important challenges for biologists is to describe and explain geographical patterns in biodiversity. Analysis of such patterns provides insight into the ecological and evolutionary processes that shape life on earth, and is also a prerequisite for conservation prioritization. Research into large-scale diversity patterns has traditionally focused on taxonomic groups for which large amounts of distributional data are available, such as birds, mammals and butterflies, with many of these systems benefitting from volunteer data collection (e.g. Robbins et al. 1989; Greatorex-Davies & Roy 2000; Newman, Buesching & Macdonald 2003). The value of volunteer schemes with regards to biodiversity monitoring has been considered for these systems, with mixed results (Lovell et al. 2009; Schmeller et al. 2009; Kremen, Ullmann & Thorp 2011).
There is an urgent need to expand the taxonomic, temporal and spatial scale of applied and theoretical biodiversity research, particularly within less accessible environments such as aquatic systems. Paradoxically, some volunteer data collection schemes have been highly successful in these environments with regards to the quantity of data collected (e.g. Pattengill-Semmens & Semmens 2003; Goffredo et al. 2010). If these data are shown to be suitable for the study of patterns of diversity, the value of such schemes will hugely increase, with implications for the collection of data in all ecosystems.
A key aspect relating to the value of volunteer data is the reliability of data returned from the protocols used to collect it. For studies performed by professional scientists, underwater visual survey protocols are often designed to minimize bias, maximize precision and ensure repeatability. Due to logistical limitations, vast sections of the world's aquatic ecosystems are rarely, or never, surveyed by professional scientists. The large pool of volunteer enthusiasts has potential to substantially augment the census capabilities of professional researchers. For example, over 8,000 surveys were performed worldwide during 2011 alone by one volunteer organization (R.E.E.F. 2012). Protocols designed for volunteers also attempt to standardize survey efforts, but must balance this requirement against the need to maintain the interest of the public. Whether data produced by such protocols are suitable for comparative studies of biological diversity remains unclear.
The development and popularity of underwater visual survey techniques using self-contained underwater breathing apparatus (SCUBA) equipment has resulted in monitoring of the underwater environment on a scale that was previously impossible. Underwater visual survey methods have been used extensively in tropical (e.g. Pattengill-Semmens & Semmens 2003) and temperate marine habitats (e.g. Goffredo et al. 2010), as well as freshwater systems (e.g. Brosse et al. 2001). Many previous studies have compared underwater visual survey protocols; most of these studies focused on identifying sources of bias within methods, often with a view to quantifying differences among protocol (e.g. Thresher & Gunn 1986; St John, Russ & Gladstone 1990; Sullivan & Chiappone 1992; Miller & Ambrose 2000; Schmitt, Sluka & Sullivan-Sealey 2002). Few, if any, of these studies have been focused on the capacity for underwater visual survey protocols to reflect actual biological patterns or to test specific ecological hypotheses. This is surprising, as it is widely acknowledged that decisions regarding the choice of methodology should be based on the study question. The likely reason for this discrepancy is that in many underwater ecosystems it is impossible to completely sample any area using any method and, without a full taxonomic list for comparison, it is difficult to quantify the performance of any particular sampling method. Our study addresses this issue by concentrating large amounts of survey effort on a very small number of sites (three) to both reliably identify any differences between two test protocols and thoroughly elucidate patterns among study sites. The techniques chosen for this study represent the most frequently used underwater visual survey methodology in published peer reviewed fish diversity studies (the belt transect) and the Roving Diver Technique (RDT) used by the Reef Environmental Education Foundation (REEF) volunteer fish survey project (Pattengill-Semmens & Semmens 2003), thought to be the largest marine species sighting database in the world, and similar to protocols used by other successful programmes. Volunteer data, such as those collected by REEF, are potentially a highly valuable resource for the marine environment, where the measurement of fundamental aspects of diversity, across expansive spatial scales, has been suggested to be a key management priority (Palumbi et al. 2008). Although studies typically vary considerably on the specific aspects of diversity they address, e.g. taxonomical relatedness (Carranza, Defeo & Arim 2011), phylogenetic diversity, functional diversity (Halpern & Floeter 2008), species diversity and community composition comparisons (i.e. α and β-diversity) are relevant to most studies and conservation objectives, and are therefore the focus of this study. As belt transects are regularly used in professional reef fish diversity studies (Kulbicki et al. 2010), they represent a logical choice with which to compare the performance of the RDT protocol. The extent to which belt transect results are consistent to those produced using RDT protocols is therefore informative regarding the utility of vast amounts of volunteer data that are currently available and collected in the future. The objective of this study is to determine whether the two protocols differ in terms of the α (i.e. within site diversity) and β-diversity (i.e. differentiation between sites) of the communities they record and in their power to detect significant differences in these biodiversity measures between these communities. We also examine how detectability (i.e. probability to detect a species that is present in a surveyed area at the time of survey) varies between protocols, as well variation associated with sites, functional groups, taxonomic groups, survey duration and underwater visibility.
- Top of page
- Materials and methods
- Supporting Information
Our results provide evidence that less standardized survey protocols used by volunteer programs may give results that are broadly consistent with those based on methods used by professional scientists. In this study the evaluated survey protocols were highly consistent with regards to comparisons of site species totals. The species richness results show differences in site ranking between the protocols, neither protocol detected any significant pairwise difference between sites. Both protocols show similar β-diversity relationships among sites, but the significance of βw values was inconsistent between the protocols. After 72 surveys per protocol, RDT surveys record significantly more species than surveys using belt transects, due to substantially higher detectability, i.e. the RDT protocol was capable of recording considerably more species per survey. Despite the observed differences in detectability between the protocols, βW analysis suggests that there is no significant difference between the protocols regarding the composition of species detected (Fig. 2).
The large number of replicates (n = 24) taken at each location in each season is at the upper end of replication for reef fish surveys, near the point at which most species capable of being detected will be observed, even in low-detection areas (MacNeil et al. 2008b). As completing more than the 24 surveys per site (per protocol) is probably beyond the reach of many survey programmes, RDT may be preferable if the research goal is to detect as many species as possible. Although previous work has suggested that both RDT and belt transects record distinct subsets of the overall species pool (Schmitt, Sluka & Sullivan-Sealey 2002), our analysis suggests that such differences may not differ from null expectations. Protocol species richness comparisons within sites A and B show that the belt transect recorded a higher diversity of species among observations (Fig. 2a), but the RDT protocol more than compensates for this by returning a larger number of observations per survey. The higher species totals of RDT survey data may be driven by factors such as time spent recording fish (Fig. 5a), area covered by the survey and/or flexibility in search methodology. Note that the actual time spent recording fish differs considerably between the two protocols (see Materials and Methods), and, as none of the sites appear to have been sampled to completion (Fig. 3), this factor may have a strong influence on this result.
Many species will be missed by both sampling protocols, and further work is required to quantify the performance of other types of underwater visual survey protocols, such as stationary counts, against those tested here. It is clear from our occupancy model results that the survey protocols tested are biased towards detecting certain functional groups (Fig. 5a.), ranging from gobies/blennies, which have extremely low detectability, to herbivores, which were observed relatively easily. Destructive methods, such as rotenone sampling, can produce more complete taxonomic inventories than underwater visual survey methods (Smith-Vaniz, Jelks & Rocha 2006); however, although these methods may be suitable for exhaustive sampling, they cannot be applied across extensive spatial scales, or within protected areas, and are therefore limited in terms of their coverage. In addition, these methods also tend to miss some species that are recorded by underwater visual survey methods (Smith-Vaniz, Jelks & Rocha 2006) and such methods may need to be combined if the study goal is to produce a full species inventory.
Species totals may be a more relevant diversity measure for conservation purposes (Gotelli & Colwell 2001) and it is common for studies to be concerned with the number of species in a given area rather than the number of species per given number of individuals (often the term species richness is used to refer to species totals). On the basis of this study, RDTs are recommended as this protocol gave results that were consistent with belt transect data while recording a larger number of species per site. A caveat to this recommendation is that relative area of survey sites should be considered as RDTs are not restricted in terms of the exact area they cover. For this study, all three sites were part of a relatively expansive coastal area and the amount of area covered was determined by dive times and swimming speeds that did not vary substantially between sites. However, if more restricted sites, such as wrecks or small reef systems, are surveyed then differences in habitat area may need to be controlled for. Further studies of this type should support our analytical approach and survey design for additional sites, including other ecosystems, such as terrestrial and freshwater communities.
With regard to species richness, differences seen between protocols at sites A and B are of concern. It is possible that species richness among observations is a poor substitute for species richness among individuals for one or both protocols. Adapting the protocol to include estimates of the exact numbers of individuals for each species may bring benefits in this regard, although any modification would also need to account for variation in detection rate among species (MacNeil et al. 2008a,b). Currently, belt transects are preferable because they return specific measures of individual abundance required for this measure of diversity. New analytical approaches also have the potential to address this issue (Yamaura et al. 2011).
The inconsistency between the protocols regarding the significance of between-site βW values is surprising given the protocol comparison results. It is possible that the constrained nature of the null models influences the probability of type II errors for presence/absence data differently to abundance data. The two βw values that did not differ significantly from null expectations in the RDT results were both close to the upper 95% confidence limit for nearly all sampling levels (Fig. 4). However, these results also suggest that belt transect data return higher βw values than the RDT data, a result that will be driven by differences in data collection protocols and is not influenced by the existence or absence of abundance data. As βw values between the two protocols were not significantly different to null expectations, this cannot be explained by the protocols detecting different species. βw values tended to decline as more surveys are completed (Fig. 4) and therefore the higher number of species returned by the RDT protocols may have resulted in lower (and possibly more accurate) βW values.
Generally speaking the RDT protocol is successful in terms of the quantity of data that have been collected. The REEF volunteer fish survey project has collected over six million sightings across over 10 000 locations and a similar program in Italy was highly successful at collecting a very large amount of data in a short period of time (Goffredo et al. 2010). The results of this study suggest that RDT protocols can be consistent with belt transects when quantifying α-diversity, providing an invaluable resource for large-scale ecological studies of biodiversity.