Fifteen Years of Rhode Island Oyster Restoration: A Performance Evaluation and Cost-Benefit Analysis

Federal, state and local non-profit organizations have long recognized the ecological and socioeconomic importance the oyster, Crassostrea virginica, represents to coastal communities. Shellfish restoration programs in Rhode Island date to the early 1900s and have been making considerable progress and gaining popularity in the past decade. To better understand both short and long term performance of oyster restoration in Rhode Island a compilation of all oyster restoration activities from 2000 to 2015 was undertaken. Restoration performance was assessed by comparisons of growth, survival, disease and recruitment over eleven years in two distinct programs; Roger Williams University’s Oyster Gardening for Restoration (2006 2014) and the North Cape Shellfish Restoration Program (2003 2008). Mean costs of restoration were weighed against cumulative value of ecosystem services provide by oyster reefs. Over 26 million oysters, encompassing 6.6 acres have been seeded in thirteen distinct restoration sites in Rhode Island waters including salt ponds, tidal creeks and open coves in Narragansett Bay. Mean growth of oysters in restoration sites was between 30-50 mm annually with mean survival of 22% and 55% for year one and two+ oysters, respectively. Mortality varies among sites and appears to be driven largely by disease. Mortality outpaces recruitment at all monitored sites leading to a decline of the population once seeding has ceased, driving the need for continued restoration to maintain desired ecosystem services. A cost-benefit model indicates Rhode Island oyster restoration is not equitable in terms of ecosystem services provided, as the cost of restoration is higher than the cumulative value of ecosystem services provided by the oyster reefs, thus, questioning the economic feasibility of restoration and emphasizing the importance of proper site selection coupled with alternate management strategies.

Mexico to the Gulf of St. Lawrence, Canada (Buroker 1983). Eastern oysters play a critical ecological role within our coastal environment. Often dubbed 'ecosystem engineers', this role has been recognized as early as Moebius's (1883) pioneering monograph on oysters and oyster culture. Oysters are capable of benthic-pelagic coupling by filtering phytoplankton and seston and transporting this organic matter to the benthos, thus supplementing benthic food webs and accelerating nutrient cycling within the system (Dame 1993, Smaal and Prins 1993, Pietros and Rice 2003. Through filter feeding activities, C. virginica increases water clarity, reduces turbidity (Cloern 1982, Newell 1988) as well as reduces carbon, nitrogen, (Hargis and Haven 1999) and pollutants from the water column . Oyster beds create complex biogenic structures, which increase species density, biomass and richness over nearby mud habitats , Manley et al. 2010, Abeels et al. 2012, Quan et al. 2012) and serve as essential fish habitat (Coen et al. 1999, Peterson et al. 2003; ultimately increasing productivity within our coastal waters (Grabowski et al. 2004, Grabowski et al. 2008).
2 Oyster beds, when healthy, can provide a direct economic benefit to coastal communities through both commercial and recreational fisheries and the infrastructure which support them. In the late 1800s and early 1900s, Narragansett Bay housed over 21,000 acres of private oyster beds resulting in annual landings of 14 million pounds (DeAlteris et al. 2000). Oyster populations are of the most degraded ecosystems in the world, with a global reef loss of 85% and reefs in New England have been considered functionally extinct (Beck et al. 2011 Monitoring restored populations and the associated habitat is a fundamental part of the restoration process and allows practitioners and managers to learn from previous efforts and progress toward more successful restoration (Brumbaugh et al. 2006 It has long been recognized the effectiveness of restoration projects must be evaluated against a reference (Fagan et al. 2008). To progress toward more effective restoration, managers must understand both, the reference of target or 'un-degraded' ecosystems as well as the reference of previous restoration methods and performance. Based on number of oysters seeded on an annual basis, Rhode Island restoration has increased by a factor of twenty in the past fifteen years. The number of restoration sites, methodology used and entities involved has also increased dramatically. The ability of oyster reefs to provide ecosystem services, including but not limited to increased water quality, habitat and fish production has gained recognition in both the scientific and political communities. Increasing water quality, habitat and fish production have been identified as federal priorities, set forth by the National Oceanic and Atmospheric Administration (NOAA).
Many approaches to achieve these goals exist including (e.g. managing combined sewage treatment outflows, restoring upland and shoreline grass habitats, implementing artificial reefs and restoring shellfish beds) and are practiced locally. Oyster restoration has been touted as a cost-effective approach Smyth 2011, Grabowski et al. 2012) and is often funded on this basis. Values of ecosystem services are likely to be highly context specific, dependent upon practice, scale of restoration, population dynamics, biophysical and chemical parameters of the given habitat and management of the restored area. Valuation of ecosystem services, in economic terms, by oyster reefs has been published in primary literature (Henderson and O'Neil 2003, Piehler and Smyth 2011, Grabowski et al. 2012 This work aims to: 1) document past oyster restoration in Rhode Island from 2003 to present, including methods and completed effort; 2) measure and analyze restoration performance and 3) incorporate these data into a costbenefit analysis based on ecosystem services provided by oysters. This information will allow us to comment on the efficacy of oyster restoration in the state and provide suggestions for future efforts. Ultimately this document will provide an additional tool for authorities to adaptively manage oyster restoration to optimize both ecological services and economic investment.

Data Compilation
To allow for analysis of restoration performance across sites, years and projects, data was compiled from multiple sources including direct field work,  Nitrogen stored in oyster shell and tissue was not accounted for, as harvest is prohibited from all restored oyster reefs in Rhode Island. Newell and Koch (2004) suggest that the oyster's ability to reduce turbidity and by depositing nutrients in biodeposits enables oyster reefs to promote the growth of submerged aquatic vegetation in shallow estuarine waters at an estimated rate of 0.005 hectare of SAV per one hectare of oyster reef. The importance of submerged aquatic vegetation as nursery ground for many coastal species is well understood (Thayer et al. 1978). Grabowski et al.   (Jones andJones 1998, Kennedy 1996), raised in a nursery grow-out for one season (June -November) and seeded on unprepared or un-cultched sites (Hancock et al. 2004(Hancock et al. , 2006(Hancock et al. , 2007DeAngelis et al. 2008). An exception to this was the season of 2008, in which oysters were set as singles and raised in an upweller for one   Table   5).
The mean sustainability index across sites was not significantly different (p=0.6022).

Cost-Benefit
Mean costs of restoration per acre of restored reef (OGRE andNorth Cape Programs, 2004 -2014) was $71,366 ± $8,592. The lack of recruitment to reefs hinders sustainability, prompting the need for maintenance seeding to preserve adequate oyster density to provide desired ecosystem services.
Mortality estimates, indicate restoration reefs need to be reseeded every six years to maintain reef integrity, causing a stepwise linear slope of cumulative restoration costs (Figure 11) targeted at 20 mm to mitigate predation pressure, however, a large variance in the size of oysters during seeding events has been observed (Hancock et al. 2004(Hancock et al. , 2006(Hancock et al. , 2007DeAngelis et al. 2008), leading to increased predator pressure on oysters which have not reached a size of predator refuge. Initial density of remote set oysters, within Rhode Island, on media (i.e. surf clam or oyster shell) is typically between 10 -200 oysters per shell leading to high inter-specific competition. After two to three years of growth post-seeding oyster density ranges between 0 and 20 oysters per shell media (Griffin, unpublished data). The precipitous drop in oyster density per shell media is largely a factor of physical space limitation. Observations of mortality in the first year post seeding also include sedimentation, as oysters can be smothered in areas of high sediment deposition and shell subsidence.
Observed first year mortality on Rhode Island oyster reefs does not appear out of the ordinary, as year one morality of 20 -30 % has been observed in other regional oyster restoration efforts (Griffin 2015).
Excluding first year mortality, highest mortality is observed in individuals with a shell length between 80 -120 mm, which is indicative of mortality caused by Perkinsus marinus. Levels of P. marinus infection build with age, as does associated percent mortality (Encomio et al. 2005), explaining mortality of the older cohorts. Survival of year 2+ oysters varied greatly between sites and within sites between years. Mortality rates of oysters can vary across space and time due to differences in habitat quality, disease and predator pressure. Part of the observed variance of mortality between years is undoubtedly due to sampling error. Mortality was based on the change of oyster density observed during annual sampling events. Oyster density on restored reefs varies greatly due to the nature of seeding, which often involves dumping totes of spat on shell off the side of boat in a predetermined area; a less than precise operation. Limited recruitment to restored reefs does not allow oyster density to become homogeneous across the site as time passes.
Haphazard quadrat sampling of reefs was employed during surveys, keeping the sample size high and consistent between years to reduce variance; however, the large standard errors associated with observed oyster densities greatly effects mortality estimates and confounds analysis comparing mortality across sites and years leading to non-significant results. Monitoring of oyster restoration efforts in the Chesapeake Bay has demonstrated year 2+ survival rates between 30 -70% (Mann and Powell 2007). We observed a mean annual survival of 55% with a range from 25% to 100%, which appears to be 35 on par with highly intensive efforts in the Chesapeake Bay.
Perkinsus marinus has been observed in eastern oysters for over 50 years along the eastern and southeastern seaboards of the United States (Smolowitz 2013). Andamari et al. (1996)  Prevalence of dermo is highly variable and directly correlates with temperature and salinity. Prevalence and intensity are generally highest in salinities greater than 12 ppt. Temperature also regulates the disease, as the prevalence and intensity oscillates with seasonal fluctuations in water temperature. Maximum prevalence and intensity generally lags 1-2 months behind maximum summer water temperature and minimum prevalence and 36 intensity lags 1-2 months behind minimum winter water temperatures (Burreson and Ragone Calvo 1996). Prevalence and intensity of dermo was similar across most sites with the exception of Smelt Brook Cove and Saugatucket River which exhibited significantly higher dermo rates compared to Potter Cove, Great Salt Pond, and Quonochontaug Pond. There is not enough variability in salinity or water temperature within the current restoration sites to influence the presence of dermo. All restoration sites with the exception of Saugatucket River experience salinities between 22 -35 ppt depending on tidal cycle and amount of precipitation. Salinity at Saugatucket River varies between 4 -24 ppt depending on tidal cycle and rainfall. The short pulses of low salinity in Saugatucket River are apparently not sufficient to extricate P. marinus, as the site has consistently high infection rates. Dermo is transmitted directly between oysters, as new infections are acquired as oysters feed and the parasite infects its host though gut epithelial tissue (Villalba et al. 2004, Bushek et al. 2002. This mechanism of transmission can cause densely populated oyster beds to be particularly susceptible to high levels of dermo. Regression analysis showed no correlation between presence of dermo and density of oysters within Rhode Island restoration sites.   (Pafford 1988). The half-life of oyster shell varies between 3 -10 years depending on the given environment (Powell et al. 2006). siltation or inadequate settlement substrate, or larval displacement greater than the study area (Dickie 1955, Hancock 1973, Wolf 1988 Morality and river flow estimates in the James River, VA have been recorded since 1994. Data shows, in years of low flow, oyster mortality rates exceed 70% and recruitment was hindered (Mann and Powell 2007). Extant subtidal oyster communities in the Chesapeake Bay are limited to upper sub-estuaries where lower salinity regimes exist (Mann and Powell 2007). Tolerated salinity ranges for oyster larval rearing is widely reported between 3 and 33 ppt (Calabrese and Davis 1970, Amemiya 1926, Carriker 1951, Davis 1958.
Optimal salinity ranges for larval rearing has been reported between 17 and 29 ppt (Calabrese and Davis 1970, MacInnes and Calabrese 1979, Amemiya 1926  Cove, but both of these sites had densities of less than 1 oyster m -2 . While this wouldn't provide much, in terms of ecosystem services, they might yet contribute to the total spawning stock biomass of oysters within our coastal waters. If we further assume that these remnant populations have been disease challenged and survived, they may contain some as yet unknown aspect of disease resistance to offer to future populations.
The cost-benefit model indicates Rhode Island oyster restoration is not equitable in terms of ecosystem services provided, as the cost of restoration is higher than the cumulative value of ecosystem services provided by the reef.
Mortality outpaces recruitment within all restoration sites, prompting the need for maintenance seeding to preserve a functioning reef in terms of ecosystem services; thus, the cost of restoration is not fixed and the cumulative cost of restoration rises at a steeper slope than the cumulative value of ecosystems services. It should be noted the ecosystem services described herein (nitrogen removal, fish production, and submerged aquatic vegetation enhancement) were calculated using data from estuaries in the southeastern United States.
Due to differences in temperature, sediment chemistry, fish assemblages and oyster productivity, we cannot assume reported values are directly comparable to oyster reefs in Rhode Island. It is, however, safe to assume reefs located in the southeastern United States perform at a higher level, in terms of production and nitrogen removal, than those found in New England waters; thus, fitting our data to this model would overestimate the ecological value of reefs. This assumption is based on warmer water temperatures in southern estuaries compared to Rhode Island, leading to a longer filtration season and higher levels of de-nitrification coupled with higher fish productivity. Grabowski et al. (2012) predict the initial investment of restoring one acre of oyster reef will be recouped through ecosystem services within 10               Figure 11. Cost-benefit model of cumulative ecosystem services provided from one acre of oyster reef versus costs of restoration. 'Actual' restoration costs represent annual operation costs from North Cape and Oyster Gardening for Restoration and Enhancement programs to maintain one acre of oyster reef. 'Theoretical' restoration costs represent a self-sustaining population after the initial seeding.
63 APENDICES Appendix A. Oyster restoration in Town Pond, Portsmouth.