INTRODUCTION

Eastern oyster (Crassostrea virginica) populations are significant components of coastal ecosystems, playing numerous important ecological, cultural, and economic roles. Eastern oyster (hereafter, “oyster”) populations often create complex reefs of living oysters of multiple size and age classes living on and among dead oyster shell material known as cultch (Fig. 1). These oyster reefs create key habitat for numerous fish, invertebrate, and bird species, many of which have large recreational and commercial value (e.g., red drum [Sciaenops ocellatus], Florida stone crab [Menippe mercenaria]) or are species of special concern (e.g., American Oystercatcher [Haematopus palliatus]). Oyster reefs also function as barrier islands in many areas, dampening wave action to reduce coastal erosion and protect human coastal communities from storm damage (Borsje et al. 2011), as well as improve coastal water quality (Coen et al. 2007). Globally oyster reef distribution has declined by as much as 85% for a variety of reasons including overharvest, disease, and poor water quality (Beck et al. 2011). Well-known U.S. oyster fisheries such as those in the Chesapeake Bay are much smaller than historic levels (Wilberg et al. 2011). The largest wild oyster fishery in the world currently is in the Gulf of Mexico, which supplies about 50% of the U.S. commercial oyster harvests (Beck et al. 2011).

Florida provides about 10% of the U.S. commercial harvest (MacKenzie 1996), with the majority of oyster landings coming from Apalachicola Bay (Dugas et al. 1997). Apalachicola Bay oysters have traditionally been viewed as a high-quality seafood product, and Apalachicola oysters are marketed by name for their size and flavor qualities. Oyster fisheries and oyster processing are a significant component of the local economy, supporting more than 1000 jobs and about half of the revenues for some coastal counties (Whitfield and Beaumariage 1977, Havens et al. 2013).

Apalachicola Bay oyster populations have a long research and management history (Fig. 2). In a report of the U.S. Commission of Fishes and Fisheries, as part of a survey of oyster regions of Apalachicola Bay, Florida, Swift (1897) said:

The oysters of this bed, especially those found near the 3-foot curve off Cat Point, are of the very finest quality, and it is probable that no better flavored oysters can be found in any part of the country. They are not only exceptionally good in flavor, but are large and fat. Swift (1897:210)

This bed [South Lump], like the others surrounding it on the north side, was formerly very productive, but it, like the others, was so overworked that it became depleted a few years ago. Since that time, these beds have been left to recuperate, and it seems probable that, if left undisturbed, they will soon recover their former productiveness. Swift (1897:203)

This is the first of several large declines in Apalachicola oyster populations reported since the late nineteenth century. The most recent occurred in 2012, when Apalachicola Bay experienced large declines in the abundance of harvestable oysters, leading the State of Florida to request a Federal Fisheries Disaster declaration from the National Marine Fisheries Service. This decline resulted in large economic losses in the region, leading to a 2012 community-based review of environmental conditions in Apalachicola Bay (Havens et al. 2013, Camp et al. 2015). This review followed decades of earlier agency and academic research on Apalachicola Bay ecology in the 1970s-2000s (Livingston 1991, 2002, 2015).

Despite the economic importance of the Apalachicola oyster fishery and expanded attention to the ecosystem services provided by oyster populations across their distribution (Coen et al. 2007, Beck et al. 2011, Seavey et al. 2011), the status of oyster populations in quantitative terms, i.e., terms useful for making management and restoration decisions, is not well known in many areas (Wilberg et al. 2011). To inform Apalachicola oyster fishery management and restoration, it is specifically critical to understand (1) the role that fishing effort has played in the current oyster fishery collapse to determine best fishing practices in the future and (2) what specific strategies or scenarios (e.g., shelling of oyster bars, restrictive harvest policies) will lead to the most rapid or most certain recovery of the fishery.

We analyzed available data in a population dynamics model to assess what mechanism(s) likely led to the collapse of the Apalachicola Bay oyster fishery. An original contribution is that our model captures the feedback between natural mortality and the accretion of shell material as substrate for oyster larvae (spat) settlement and growth, and the linkages between harvest (which removes both oyster shell material and live oysters) and recruitment. We then assessed the efficacy of alternative management strategies (e.g., habitat restoration, fishery closures) and scenarios (area for and frequency of adding shell material) to accelerate oyster population and fishery recovery to help inform planning efforts for community-led restoration programs designed to promote resilience in this resource-dependent community (Camp et al. 2015).

Note that we did not study or reach any conclusions about any effect of water withdrawals affecting the Apalachicola River Basin or oyster populations in Apalachicola Bay. This is an area that warrants future research.

STUDY SITE

Apalachicola Bay is a shallow estuary (mean depth <3 m) of approximately 63,000 ha enclosed by a series of barrier islands with an east-west orientation. Geologic surveys of the bay suggest that the primary oyster bars are perpendicular to the orientation of the bay along ancient sandy deltas, and these bars became expansive 1200-2400 years ago (Twichell et al. 2010). The primary source of freshwater input into Apalachicola Bay is the Apalachicola River, and river discharge has a strong influence on the salinity, nutrient dynamics, and other aspects of the Apalachicola Bay ecosystem (Livingston et al. 1997).

Fishery overview

The commercial oyster fishery in Apalachicola was first described in the 1880s, and extensive surveys in the 1890s reported established fisheries with commercial canning operations, as well as documentation of oyster bars being “overworked” and no longer producing a commercial harvest (Swift 1897, Dugas et al. 1997; Fig. 2). Oyster landings from Apalachicola Bay in the last half century average between 91,000 and 272,000 kg of meat, or about 90% of Florida’s commercial oyster harvest (Dugas et al. 1997; Fig. 3), with the majority of harvest coming from public reefs where oysters are harvested via hand tonging (Whitfield and Beaumariage 1977). As early as 1881, it was recognized by Florida statute that the recycling or placing of oyster shell on oyster reefs to provide substrate for oyster spat (known as “shelling”) was important to promote sustainable oyster harvest (Whitfield and Beaumariage 1977). In 1949, a management program was established to replace oyster shell on public reefs, and the amount of material has varied annually depending on funding and availability of material (Whitfield and Beaumariage 1977).

In 1985, Hurricane Elena caused significant damage to oyster resources in Apalachicola Bay, leading to highly restrictive regulations, on-water harvest check stations, and intensive shelling operations on a subset of reefs (Berrigan 1990). Beginning in 1986, a revised landings and effort reporting system was required for all commercially harvested marine species, in contrast to the prior voluntary reporting program in place. Based on data since 1986, the number of Apalachicola Bay oyster harvesters declined from about 1000 in the late 1980s to around 400-600 throughout most of the 1990s and early 2000s, before increasing since 2008 to about 1000 license holders at present (Fig. 3a). The number of oyster fishing trips follows a similar pattern, with about 30,000 trips reported in 1988, declining to about 10,000 trips in the mid-1990s, and then varying between 10,000 and 25,000 trips until 2006, when the number of trips increased to about 40,000 annually in recent years. Large declines in landings were reported beginning in the fall of 2012, and landings and trips declined dramatically in 2013 (Camp et al. 2015). Oyster regulations in Apalachicola Bay are currently managed using a system of seasons, spatial closures, bag limits, and size limits, but on-water check stations and a bag tax to fund research and monitoring programs were ended in the early 1990s (Fig. 2).

METHODS

Oyster stock assessment model

We developed an age-structured oyster stock assessment model that reconstructs historical abundance patterns and allows for exploration of future alternative management options. The model represents a single oyster population in some area of interest; that region may be some large management area like Apalachicola Bay or some much smaller habitat type or site within a larger region. The model is implemented in an Excel spreadsheet to allow portability and ease of examination of model structure and calculations; an example copy of this spreadsheet is available as a supplemental file (see Appendix 1).

Model population dynamics calculations (growth, survival, recruitment) are made at a monthly time resolution to account for the rapid growth and mortality of oysters, to assist in interpretation of seasonal harvesting data, and for evaluation of seasonal harvesting policies, e.g., seasonal closures. Three time-accounting variables are used in the model equations: y for calendar year (y = 1,...,n_y), m for month of year (m = 1,...,12), and t for month from start of a time simulation (t = 1,...,12n_y). The model predicts matrices of oyster numbers N_a,t and shell lengths L_a,t by month of age. Using monthly fishing efforts E_y,m, the model predicts monthly catches C_y,m for statistical comparison to historic data, while allowing for interannual variations in recruitment, growth, and survival (equation 1 in Table A2.1, Appendix 2). For our Apalachicola Bay case history, observed fishery landings and effort data by month were compiled for 1986-2013 from information provided by the Florida Fish and Wildlife Conservation Commission. We developed a standardized index of oyster recruitment using fisheries-independent survey data of oysters by 5-mm size classes collected by the Florida Department of Agriculture and Consumer Services (DACS) on the major commercial fishing reefs in Apalachicola Bay (available from 1990 through 2013). Full details on our population dynamics model and its application to the Apalachicola Bay case study are available in Appendix 2.

Simulation of management actions

When tuned to the uncertainty in oyster population dynamics in Apalachicola Bay, this model can be used to evaluate future fishery outcomes of alternative management actions. As an example, how harvest or environmental perturbations affect persistence of shell material essential for successful recruitment represents a key uncertainty in managing oyster populations. If shell material is removed from the oyster bars as fishermen cull legal oysters from sublegal sizes and associated shell material, and these sublegal oysters and shell are discarded away from the oyster bar, then this loss rate (discard mortality of both live sublegal oysters and shell material) could be substantial. Such evaluations are a critical part of any adaptive management program designed to learn more about the system (Camp et al. 2015).

We used the model to assess the effect on future oyster fishery landings of a variety of potential management actions under different assumptions about oyster recruitment patterns and processes. We first assumed future average oyster recruitment levels similar to those observed in 2004-2013 and explored four management action scenarios: (1) no management action taken; (2) no management action taken, but assuming a different function form of the recruitment relationship (i.e., assuming a Beverton-Holt recruitment function rather than a Ricker recruitment function); (3) initiating a substantial shell addition program involving restoration of about 50 ha (a little more than the historic average annual shelling) for each of four years (2014-2018); and (4) reducing fishing effort by half (from around 4000 trips/month to 2000) over the next six years (2014-2020). We then evaluated actions 1, 3, and 4 under the alternative assumption that future average oyster recruitment remains low (similar to 2011-2012 levels). Finally, we assessed how a future 20% annual loss rate of shell material would impact oyster population recovery under no action (scenario 1) and reshelling (scenario 3).

RESULTS

Oyster stock assessment model

The oyster stock assessment model (Appendix 2) appears to represent well the oyster population dynamics in Apalachicola Bay and results in a remarkably good statistical explanation of historical catches and major trends in fisheries-independent survey data. Our assessment results suggest that the Apalachicola Bay oyster population is probably not recruitment overfished; i.e., the observed low recruitments are not anticipated to be strictly because of overharvest of adults. Rather, the 2012 collapse instead was likely driven by lower-than-average abundance or survival of sublegal (juvenile) oysters in the years preceding the collapse. This reduction in recruitment not only reduced the biomass of oysters available to harvest, but from a population resilience perspective, likely reduced the amount of dead shell material available as area for larval settlement.

An important and surprising result of our work was the low estimate of the area of oyster bar needed to produce the estimated abundance of legal-size oysters, in other words, to support the observed harvests (A_total, about 500 ha). The low A_total estimates would imply very high monthly exploitation rates (U_t), in the range of 0.1-0.15 for recent years. This is much higher than U_t estimated from about 1995-2010 of around 0.05 per month, but similar to the estimates from about 1988-1990 (Fig. 4). This low estimate of A_total has multiple potential interpretations: (1) the Apalachicola Bay fishery is being supported by recruits from a very small but productive proportion of the total oyster bar area (total area of oyster habitat including subtidal areas estimated to be about 4800 ha); (2) fishery catches have been grossly under-reported; (3) DACS harvest data are not representative of average oyster densities; (4) the impact of a unit of fishing effort (fishery catchability, q, and hence U_t) has been greatly overestimated; (5) the total oyster bar area has been overestimated based on existing geological (bottom type) and other survey information; and/or (6) the model allows upward bias in exploitation rate estimates by not properly accounting for erosion of size structure at age (selective removal of faster growing individuals).

The basic problem is most likely that the model predictions of sublegal and legal abundances, i.e., size distribution, are not in fact reasonable for such high exploitation rates (interpretation 6). If exploitation rates were in reality as high as predicted, most legal oysters would be removed within a few months of reaching the 76.2-mm legal length, and the size distribution would thus be far more severely truncated than observed; i.e., the legal/sublegal density ratio would be much lower. Instead, the model “allows” very high exploitation rates without erosion in the predicted size structure because it assumes regeneration of the length distributions of sizes each month. The only way to avoid this faulty regeneration assumption would be to use a much more complex model structure involving growth-type groups. However, when we developed growth-type-group models based on five-year periods of time (Appendix 2), these models resulted in similar estimates of A_total of around 500 ha, prohibiting a conclusive dismissal of the small A_total estimate as an artifact of the model structure.

The lack of clarity regarding the area of oyster production, A_total, reverberates throughout the model results. If A_total is larger than estimated, as implied by the above information, our assessment suggests a large, relatively unproductive oyster population scarcely affected by fishing harvest. Alternatively, if A_total is truly small, our assessment would suggest a highly productive population that has been relatively heavily impacted by recent fishing harvest. Therefore, although our best assessments of the available data suggest that the collapse of the Apalachicola Bay oyster fishery was not strictly due to overfishing, the model was unfortunately unable to clearly resolve the historical role that harvest has played or the relative productivity of the oyster population in Apalachicola Bay.

Simulation of management actions

The model predicted that under average “normal” recruitment and mortality rates observed from 2004-2013, oyster populations would recover in 5-10 years, even with no management action (blue lines, Fig. 5a). If recruitment remains equal to the 2004-2013 average, then adding shelling of 50 ha per year for the next 4 years was predicted to increase oyster yields to 2008-2010 levels in about 5 years, even if fishery effort remains high (green lines, Fig. 5a). Interestingly, shelling only provided a small reduction in recovery time compared with the “no action” scenario. With the same assumption of recruitment equal to the 2008-2012 average, reducing effort by half over the next 6 years (2014-2020) would result in increasing yields within 5 years, but such a reduction in effort would obviously reduce overall yield (red lines, Fig. 5a) and likely have deleterious economic effects on the community. The effects of closures or shelling additions were small relative to the effect of assuming a different functional form for the stock recruitment relationship (gray lines, Fig. 5a).

When we ran the model under the assumption that the low oyster recruitment observed in Apalachicola Bay during 2012-2013 would continue into the future, it predicted that oyster fishery recovery was less likely to occur in the absence of management actions (blue lines, Fig. 5b). If recruitment remains at 2012-2013 levels, our model predicted that even with shelling 50 ha per year for four years, the oyster population would continue to decline in 2014-2020 (red lines, Fig. 5b). Even with very high shelling rates (162 ha per year), the low-recruitment oyster population was predicted to increase only slightly (green lines, Fig. 5b) and only during the four years when shell additions take place (green lines, Fig. 5b).

A central finding is the importance of the protection of shell habitat and the maintenance of a “positive” shell budget. We found that the oyster fishery is unlikely to recover if shell loss, e.g., from storms, culling practices, or ocean acidification, is higher than shell deposition from natural or restoration actions (i.e., if the shell budget is negative). At a 20% annual shell loss rate, our model estimated relatively rapid oyster fishery collapse (red lines, Fig. 5c) compared with the recovery predicted even without management actions (blue lines, Fig. 5c) if no shell loss is assumed.

Overall, the key result from our simulations was identifying the pivotal role that recruitment rates likely play in oyster population recovery. If recruitment levels return to the average observed in 2004-2013, then oyster yields are likely to recover in about five years without any management action. If recruitment rates remain low (similar to 2012-2013), then the likelihood is high for a very slow oyster population recovery or even collapse. There are multiple factors that could lead to low recruitment at present or in the future, including:

1. Changes in A_total caused by environmental factors and/or fishing. In this scenario, declines in the total area of oyster habitat would result in overestimates in oyster population levels based on extrapolation from DACS surveys from a small subset of oyster bars.
2. Removal of dead shell during the harvesting process, leading to a negative shell budget (shell losses exceeding deposition) and an overall decline in settlement area.
3. Irreversible increases in estimated relative natural mortality rate (M) because of invasion or expansion in the system of oyster predators or diseases.
4. Nonstationarity in the stock-recruit relationship, meaning that uncorrelated random deviations from the mean stock-recruitment relationship can lead to very misleading assessments, while long-term recruitment trends because of habitat changes (i.e., settlement area, larval food supply) are masked. In traditional fisheries stock assessments, we sometimes attempt to deal with this structural problem by using virtual population analysis methods to back-calculate recruitment without assuming any underlying stock-recruitment relationship; however, we were unable to determine parameters for a virtual population analysis model with oyster populations because we did not know the age structure of the harvested population.

DISCUSSION

Factors contributing to the 2012 oyster population collapse based on available data

What led to the oyster population collapse in Apalachicola Bay in 2012-2013? Our results suggest that the 2012-2013 Apalachicola oyster population collapse was likely due to low recruitment and/or low sublegal survival rates. Our results warn that this decline may have resulted in or resulted from a decrease in larval settlement area (dead shell), which could severely retard population recovery or even send the stock into irreversible decline, depending on future recruitment and shell dynamics. Although the Apalachicola Bay oyster fishery has proven resilient over its >150-year history, this fishery now may be at a crossroads in terms of continued existence, and if recruitment levels remain low, then large-scale restoration programs may be necessary to avoid an irreversible collapse.

Perhaps the most important finding from our work is that none of the available data give superior estimates of historical exploitation rates fishing impacts. The sudden decline in Apalachicola Bay oyster landings in 2012 was preceded by several years of increasing harvest and effort. We initially suspected a case of overfishing led to the collapse of the oyster fishery; however, our analyses suggest a much more complex but classic problem in fish stock assessment: we can generally attribute the observed changes in relative abundance either to fishing or productivity changes, but we can never be sure which was more important without good independent data on absolute exploitation rates over time. With the data currently available for Apalachicola Bay, we cannot be sure whether we are dealing with a small oyster population that has been subject to strong fishing impacts or a larger population that has been subject to strong environmental influences that have impacted the long-term carrying capacity. In the latter case, the population may recover, but, if the long-term carrying capacity is reduced, it may not recover to the same historic levels.

Factors likely not contributing to the 2012 oyster population collapse based on available data

Our results are notable for what they did not find. Within the Apalachicola community there are two widespread hypotheses related to driving forces of the oyster fishery collapse. First, there is widespread concern that the oyster population collapse in 2012 was related to the Deepwater Horizon oil spill that occurred in March 2010. In a related project, a large number of sediment, water, and animal tissue samples were collected in 2012 by the University of Florida and no pollutants were detected (Havens et al. 2013, Camp et al. 2015). This corroborates results from sampling by state and federal agencies immediately after and in the years following the oil spill. Our model results suggest that the decline in sublegal oyster abundance in Apalachicola Bay did not begin until 2012, two years after the oil spill. To the best of our knowledge, the Apalachicola Bay oyster population was not directly impacted by oil or oil dispersants used during the 2010 Deepwater Horizon oil spill.

A second recent concern among the Apalachicola oyster fishing community and resource managers is the impact of low freshwater inputs into Apalachicola Bay from drought conditions within the Apalachicola-Chattahoochee-Flint basin. During 2011-2012, the Apalachicola-Chattahoochee-Flint basin experienced extensive drought (Palmer Drought Severity index of severe to extreme, https://www.drought.gov/drought/regional-programs/acfrb/acfrb-home), leading to low freshwater discharge into Apalachicola Bay and higher than normal salinity measures across several of the historically important oyster harvesting reefs (Havens et al. 2013). A series of previous studies have noted positive correlation between high-salinity drought conditions and oyster disease-related mortality (Petes et al. 2012), as well as complex relationships between estuarine freshwater discharge and oyster harvest (Wilber 1992, Turner 2006, Livingston 2015). We did not find correlations between Apalachicola River discharge measures (average monthly, total annual, total monthly, or coefficient of variation on annual discharge, mean seasonal, or total seasonal) and our estimated relative natural mortality rate (M) or oyster recruitment rates (example Fig. 6). The overall relationships between freshwater flows, drought frequency and severity, oyster recruitment, and harvest dynamics remain unclear, and this is an area of ongoing work.

Management implications: oyster stock rebuilding scenarios

A key finding from simulations of management scenarios is that oyster recruitment likely drives the Apalachicola Bay oyster fishery. We are uncertain as to the extent oyster recruitment can be influenced by management actions. This is seen in the relatively minor effect on fishery recovery time from effort reductions or shelling compared with the effect of differing recruitment averages. This uncertainty is largely because of a lack of understanding of the functional relationships between shelling and recruitment, and the inability of our assessment model to clearly define recent oyster harvest rates.

Although our model results suggest that the addition of shell as cultch material will have a minor effect on fishery recovery time under average recruitment levels, we conclude that shell addition may still be the best management action to follow. Shelling should not be expected to guarantee recovery of the fishery; but shell addition would be expected to reduce the risk of stock collapse in case low recruitment continues, while allowing for continued harvests. Adding shell material as cultch would be a critical management action if the current shell budget is deficient, with shell losses on existing reefs (from harvest-related discard, environmental disturbances, or other factors) exceeding deposition of new shell from the natural mortality of oysters. Although shell additions may restore settlement sites for oyster recruits, if oyster recruitment remains low for other reasons, then even large amounts of shelling may not lead to rapid oyster population recovery. However, based on our assessment model, shell addition is likely a better management action to achieve stated goals of oyster population and fishery recovery than fishery reductions or closures. Overall our results suggest that shell habitat as cultch should be as carefully managed as the oyster fishery for live oysters and that oyster shell material, in terms of available area, recruitment of new shell, and prevention of shell removal by harvest or storms, should be quantified and tracked.

Although recovery of the Apalachicola Bay oyster fishery may be possible without management actions, this recovery depends upon uncertain community dynamics. One uncertainty is how the fishing industry will respond over this recovery period. If the legal oyster biomass remains low, it is possible that fishers will not be willing to exert high fishing effort given costs associated with a day of fishing and low potential harvests. Such declines (voluntary or otherwise) in effort could hasten recovery; however, oyster prices are currently at near-record high levels, and alternative employment opportunities remain limited in this area. If prices and fishing effort remain high and increase, respectively, the chance of recovery could decrease as recovering oyster stocks are rapidly removed by harvest.

Management implications: policy and research

Our study suggests at least two policy issues that need to be addressed. First, it suggests that intense shelling (162 ha) should be immediately undertaken each year to counter the risk of irreversible fishery collapse, although we are uncertain at what density the shelling should occur to promote larval settlement and persistence of reef material. In 1986-1987, following 1985’s Hurricane Elena, about 156 ha of oyster reef in Apalachicola Bay was successfully restored through a combination of restrictive harvest and shelling at a density of about 472 m³ of shell per hectare (Berrigan 1990). Within 18 months of completing this restoration, these oyster bars supported 587 oysters per m² and more than 22 oysters per m² of legal size (76.2 mm), leading Berrigan (1990) to conclude that the restoration costs were recovered after one harvest season from this area. Most shelling efforts in Apalachicola Bay have been at various densities and much smaller in scale, usually averaging about 40 ha in size, and these restoration efforts have varied widely in area and frequency in the last 20 years (Fig. 3e). Natural shell deposition occurs following oyster mortality events, which may help to partially explain the rapid oyster recovery in Apalachicola Bay following Hurricane Elena. Other than the Berrigan (1990) study, no oyster restoration efforts have been rigorously evaluated to inform future strategies, such as understanding the trade-off between shelling a small area at a high shell density versus shelling a larger area at a lower shell density. Because restoration costs have increased greatly since Berrigan (1990), this type of information is critical to informing effective restoration projects.

Second, actions (including shelling) should be designed to provide more opportunity for learning about the system (Camp et al. 2015). Uncertainty in shelling density is precisely linked to uncertainty in the total productive area of the current oyster fishery. Swift (1898) estimated the total area of commercially viable oyster bar in Apalachicola Bay to be about 4942 ha, whereas Rockwood et al. (1973, as cited in Whitfield and Beaumariage 1977) estimated about 2023-2428 ha in Apalachicola Bay in the early 1970s. More recent surveys of Apalachicola Bay (Twichell et al. 2010) have focused on the geologic features that support oyster reefs and do not provide density estimates comparable to those of earlier surveys. We assessed available geographic information system layers and estimated that the existing oyster reef area was about 4000 ha, but we do not have a good understanding of how the DACS oyster survey data apply to this area; i.e., we cannot extrapolate density estimates from DACS survey data to this entire area. Determining whether the total area (A_total) of commercially viable oysters has changed in Apalachicola Bay is a key area for future work.

It is also uncertain what the impacts of fishery practices are on the persistence of shell material as cultch and sublegal oysters on oyster reefs. Swift (1897) warned of this potential for loss of shell material in his surveys of the Apalachicola Bay oyster fishery and suggested that “it is doubtful whether the law regarding the taking of small oysters and the culling of the oysters, especially the latter, are strictly complied with by the oystermen, yet it is of the greatest importance that they should realize that this law should be strictly obeyed if they wish to maintain the productiveness of the beds and thus insure themselves a livelihood in the future.” Whether culling and discarding currently take place on the bars where the tonging occurs or in areas off of the bar is not known, but it is a sensible practice to only cull in the same location that tonging occurs. The key existing uncertainties in informing management actions regarding shelling density, productive fishery area, and availability and persistence of cultch material, as well as current harvest rates and effects of changing environmental conditions, disease, and oyster predator responses, all likely influence oyster recruitment levels that apparently drive the fishery. These uncertainties can be addressed, and their reduction is critical to informing decision making and bolstering the resilience of the Apalachicola Bay oyster fishery (Camp et al. 2015).

CONCLUSION

The Apalachicola Bay oyster fishery is currently at the lowest level observed in recent decades; however, this is not the first oyster population collapse in Apalachicola Bay (Swift 1898, Andree 1983, Berrigan 1990, Havens et al. 2013). For more than 120 years, various reports and workshops have repeated the same key uncertainties for oyster resources in Apalachicola Bay (i.e., unknown total area of oysters, unknown shell budget dynamics) and made the same types of management recommendations to address fundamental fishery practices, such as preservation and cultivation of shell substrate, seasonal closures, and size limits to protect oyster bars from being “overworked” (Table A2.4, Appendix 2). These recommendations often highlight recurring issues related to poor compliance with existing regulations, including high harvest rates of undersized oysters, harvest from closed areas, or culling and discarding that occur off of oyster bars. The results of our stock assessment model suggest that Apalachicola Bay oyster resources will respond positively to management actions, particularly actions that improve availability and area of shell substrate. Our simulation results suggest that if recruitment does not return to some long-term stationary level similar to past averages or if the resilience of the Apalachicola Bay oyster resource changes over time (Camp et al. 2015), then these types of management actions may not be effective.

A variety of state and federal restoration programs totaling more than U.S. $10 million are currently committed to the Apalachicola oyster fishery and community. The oyster industry is likely to play a large role in determining how these funds are spent. The potential exists for restoration to be effective, given the success of prior oyster restoration efforts coupled with intensive fishery management for oyster resources in Apalachicola Bay (Berrigan 1990). Although restoration and management strategies are known, whether or not to follow these practices and how to use available restoration funds are choices to be made by the local community that are likely to determine the long-term viability of the Apalachicola Bay oyster fishery.

LITERATURE CITED

Andree, S., editor. 1983. Apalachicola oyster industry: conference proceedings. Report No. 57. University of Florida Sea Grant Program, Gainesville, Florida, USA.

Beck, M. W., R. D. Brumbaugh, L. Airoldi, A. Carranza, L. D. Coen, C. Crawford, O. Defeo, G. J. Edgar, B. Hancock, M. C. Kay, H. S. Lenihan, M. W. Luckenbach, C. L. Toropova, G. F. Zhang, and X. M. Guo. 2011. Oyster reefs at risk and recommendations for conservation, restoration, and management. BioScience 61:107-116. http://dx.doi.org/10.1525/bio.2011.61.2.5

Berrigan, M. E. 1990. Biological and economical assessment of an oyster resource development project in Apalachicola Bay, Florida. Journal of Shellfish Research 9:149-158.

Borsje, B. W., B. K. van Wesenbeeck, F. Dekker, P. Paalvast, T. J. Bouma, M. M. van Katwijk, and M. B. de Vries. 2011. How ecological engineering can serve in coastal protection. Ecological Engineering 37:113-122. http://dx.doi.org/10.1016/j.ecoleng.2010.11.027

Camp, E. V., W. E. Pine III, K. Havens, A. S. Kane, C. J. Walters, T. Irani, A. B. Lindsey, and J. G. Morris. 2015. Collapse of a historic oyster fishery: diagnosing causes and identifying paths toward increased resilience. Ecology and Society 20(3):45. http://dx.doi.org/10.5751/ES-07821-200345

Coen, L. D., R. D. Brumbaugh, D. Bushek, R. Grizzle, M. W. Luckenbach, M. H. Posey, S. P. Powers, and S. G. Tolley. 2007. Ecosystem services related to oyster restoration. Marine Ecology Progress Series 341:303-307. http://dx.doi.org/10.3354/meps341303

Dugas, R. J., E. A. Joyce, and M. E. Berrigan. 1997. History and status of the oyster, Crassostrea virginica, and other molluscan fisheries of the U.S. Gulf of Mexico. Pages 187-210 in C. L. MacKenzie, Jr., V. G. Burrell, Jr., A. Rosenfield, and W. L. Hobart, editors. The history, present condition, and future of the molluscan fisheries of North and Central America and Europe. Volume 1: Atlantic and Gulf Coasts. NOAA Technical Report NMFS 127. U.S. Department of Commerce, Woods Hole, Massachusetts, USA.

Havens, K., M. Allen, E. Camp, T. Irani, A. Lindsey, J. G. Morris, A. Kane, D. Kimbro, S. Otwell, B. Pine, and C. Walters. 2013. Apalachicola Bay oyster situation report. Technical Publication 201. University of Florida Sea Grant, Gainesville, Florida, USA.

Hayes, P. F., and R. W. Menzel. 1981. The reproductive cycle of early setting Crassostrea virginica (Gmelin) in the northern Gulf of Mexico, and its implications for population recruitment. Biological Bulletin 160:80-88.

Livingston, R. J. 1991. Historical relationships between research and resource management in the Apalachicola River estuary. Ecological Applications 1:361-382. http://dx.doi.org/10.2307/1941897

Livingston, R. J. 2002. Trophic organization in coastal systems. CRC, Boca Raton, Florida, USA. http://dx.doi.org/10.1201/9781420040852

Livingston, R. J. 2015. Climate change and coastal ecosystems: long-term effects of climate and nutrient loading on trophic organization. CRC, Boca Raton, Florida, USA.

Livingston, R. J., X. Niu, F. G. Lewis, III, and G. C. Woodsum. 1997. Freshwater input to a gulf estuary: long-term control of trophic organization. Ecological Applications 7:277-299. http://dx.doi.org/10.1890/1051-0761(1997)007[0277:FITAGE]2.0.CO;2

MacKenzie, C. L. 1996. History of oystering in the United States and Canada, featuring the eight greatest oyster estuaries. Marine Fisheries Review 58:1-78.

Petes, L. E., A. J. Brown, and C. R. Knight. 2012. Impacts of upstream drought and water withdrawals on the health and survival of downstream estuarine oyster populations. Ecology and Evolution 2:1712-1724. http://dx.doi.org/10.1002/ece3.291

Seavey, J. R., W. E. Pine III, P. Frederick, L. Sturmer, and M. Berrigan. 2011. Decadal changes in oyster reefs in the Big Bend of Florida’s Gulf Coast. Ecosphere 2:art114. http://dx.doi.org/10.1890/es11-00205.1

Swift, F. 1897. Survey of the oyster regions of St. Vincent Sound, Apalachicola Bay, and St. George Sound, Florida. U.S. Commission of Fishes and Fisheries, Extracted from Report of Commissioner for 1896. U.S. Commission of Fishes and Fisheries, Washington, D.C., USA.

Swift, F. 1898. The oyster-grounds of the west Florida coast: their extent, conditions, and peculiarities. Pages 185-187 in Proceedings and Papers of the National Fishery Congress, Tampa, Florida, January 19-24, 1898. U.S. Commission of Fish and Fisheries, Washington, D.C., USA.

Turner, R. E. 2006. Will lowering estuarine salinity increase Gulf of Mexico oyster landings? Estuaries and Coasts 29(3):345-352. http://dx.doi.org/10.1007/bf02784984

Twichell, D., L. Edmiston, B. Andrews, W. Stevenson, J. Donoghue, R. Poore, and L. Osterman. 2010. Geologic controls on the recent evolution of oyster reefs in Apalachicola Bay and St. George Sound, Florida. Estuarine, Coastal and Shelf Science 88:385-394. http://dx.doi.org/10.1016/j.ecss.2010.04.019

Whitfield, W. K., Jr., and D. S. Beaumariage. 1977. Shellfish management in Apalachicola Bay: past-present-future. Pages 130-140 in R. J. Livingston and E. A. Joyce, Jr., editors. Proceedings of the Conference on the Apalachicola Drainage System. Florida Marine Resource Publication No. 26. Florida Department of Natural Resources, Tallahassee, Florida, USA.

Wilber, D. H. 1992. Associations between freshwater inflows and oyster productivity in Apalachicola Bay, Florida. Estuarine, Coastal and Shelf Science 35:179-190. http://dx.doi.org/10.1016/s0272-7714(05)80112-x

Wilberg, M. J., M. E. Livings, J. S. Barkman, B. T. Morris, and J. M. Robinson. 2011. Overfishing, disease, habitat loss, and potential extirpation of oysters in upper Chesapeake Bay. Marine Ecology Progress Series 436:131-144. http://dx.doi.org/10.3354/meps09161