Classic paradigms in a novel environment: inserting food-web and productivity lessons from rocky...

Classic paradigms in a novel environment: insertingfood web and productivity lessons from rocky shoresand saltmarshes into biogenic reef restoration

F. Joel Fodrie1*, Antonio B. Rodriguez1, Christopher J. Baillie2, Michelle C. Brodeur1,

Sara E. Coleman1, Rachel K. Gittman1, Danielle A. Keller1, Matthew D. Kenworthy1,

Abigail K. Poray1, Justin T. Ridge1, Ethan J. Theuerkauf1 and Niels. L. Lindquist1

1Institute of Marine Sciences, University of North Carolina at Chapel Hill, 3431 Arendell Street, Morehead City,

NC 28557, USA; and 2Marine Science Center, Northeastern University, 430 Nahant Road, Nahant, MA 01908,

USA

Summary

1. Gradients in competition and predation that regulate communities should guide biogenichabitat restoration, while restoration ecology provides opportunities to address fundamental

questions regarding food web dynamics via large-scale field manipulations.2. We restored oyster reefs across an aerial exposure gradient (shallow-subtidal-to-mid-intertidal) to explore how vertical gradients in natural settlement, growth and interspe-

cific interactions affected the trajectory of man-made shellfish reefs.3. We recorded nearly an order-of-magnitude higher oyster settlement on the deepest (subtid-al) reefs, but within a year abundance patterns reversed, and oyster densities were ultimately

highest on the shallowest (intertidal) reefs by over an order-of-magnitude.4. This reversal was due to (i) significantly elevated survivorship on intertidal reefs and

(ii) larger surviving oysters on intertidal reefs. These patterns are likely to have developedfrom greater levels of biofouling and predator abundance (e.g. stone crabs, gastropods) on

deeper reefs where aerial exposure was <5% of the monthly tidal cycle.5. Synthesis and applications. The success of restoration initiatives involving habitat-formingspecies can be enhanced by accounting for the biotic interactions that regulate population fit-

ness. In littoral systems, vertical gradients in predation, competition and disturbance can beexploited to guide restoration of vegetated (e.g. mangrove, seagrass) or biogenic reef habitats.

In particular, our results demonstrate that paradigms of vertical zonation learned from therocky intertidal and saltmarshes also describe the fate of restored shellfish reefs. As with

rocky shores, the lower vertical limit of adult oyster distribution in our study system wasmost likely driven by predatory and competitive (i.e. smothering) interactions, with a thresh-old depth at c. 5% daily aerial exposure. Below this depth, experimentally restored reefs

failed completely. As with Spartina saltmarsh, accumulation of oyster biomass was greatest atan intermediate vertical position relative to mean sea level (i.e. mid-to-low intertidal). Our

developing model proscribes a vertical ‘hot spot’ for restoration efforts to maximize biogenicreef fitness and production.

Key-words: competition, growth, inundation–productivity gradient, landscape ecology, pre-dation, restoration ecology, rocky intertidal ecology, shellfish reefs, vertical zonation

Introduction

Within littoral systems, a number of fundamental ecologi-

cal principles related to the vertical zonation of species

could be tested as guides for biogenic habitat restoration.

For example, classical experimental work along rocky

shores has shown that interacting forces such as competi-

tion (Connell 1961), predation (Paine 1966) and distur-

bance (Dayton 1971) set the lower vertical distribution

limit for many species and also maintain overall diversity

patterns across depths. Thus, many species are restricted

to discrete zones of the intertidal based on their respective*Correspondence author. E-mail: [email protected]

© 2014 The Authors. Journal of Applied Ecology © 2014 British Ecological Society

Journal of Applied Ecology 2014, 51, 1314–1325 doi: 10.1111/1365-2664.12276

abilities to withstand abiotic stresses associated with aerial

exposure better than their ‘enemies’ (Wethey 1984). While

these findings have not been directly applied to guide the

restoration of intertidal habitats such as mangroves, seag-

rasses or shellfish reefs, there is evidence that these para-

digms regulate community structure across diverse taxa

and habitat types. In coastal saltmarshes, for instance,

competition and disturbance also regulate the vertical

(upper) limits of species’ distributions (reviewed in

Bertness & Silliman 2014), with species-specific patterns of

plant biomass also affected by consumer pressure and

interspecific facilitation (Gittman & Keller 2013). More-

over, empirical data from saltmarshes show that the verti-

cal range (Redfield 1972) and productivity (Morris et al.

2002) of a dominant macrophyte, Spartina alterniflora, are

tightly linked to tidal amplitude and inundation period,

which dictate the biotic and abiotic constraints for this

habitat-forming plant. With these concepts in mind, we

investigated the degree to which restored eastern oyster

Crassostrea virginica reefs – another biogenic habitat-

forming species – may be regulated by mechanisms

extrapolated from rocky and marshy littoral systems.

Restoration science has moved in to the limelight of

mainstream ecology over the last 25 years as efforts to

mitigate or reverse global habitat degradation have

become increasingly urgent (Ormerod 2003). When

applied in conjunction with a priori hypothesis testing,

restoration ecology as a field promotes an improved

understanding of nature via a powerful and easily recog-

nizable experimentalist’s tool: large-scale field manipula-

tions of habitats or populations. The recent elevation of

ecosystem restoration as a discipline, however, has been

defined by practice often racing ahead of theory (Peterson

& Lipcius 2003). Without the underpinnings of population

and community ecology, many restoration plans have

been plagued by trial-and-error approaches, followed by

ad hoc triage to determine what mechanisms drove project

success or failure (or no analyses in unreplicated designs).

Subsequently, these efforts have not provided useful feed-

back to inform future restoration strategies.

Coastal marine ecosystems are particularly relevant for

considering the benefits and efficacy of restoration. Mosa-

ics of saltmarshes, seagrasses, mangroves, mudflats and

shellfish reefs provide globally important ecosystem goods

and services in terms of water filtration, fishery produc-

tion, shoreline stabilization and climate buffering (e.g.

Grabowski et al. 2012). Due to accelerating human devel-

opment concentrated along coastal margins, these habitats

are also among the most threatened world-wide, with

alarming losses documented for submerged and wetland

vegetation (30–50%; Lotze et al. 2006), and shellfish reefs

(60–85%; Zu Ermgassen et al. 2012).

Although shellfish reefs are highly threatened and cost-

competitive to restore, the amount of man-made shellfish

habitat lags one-to-two orders-of-magnitude behind the

spatial extent of restoration of other biogenic coastal hab-

itats (Grabowski et al. 2012). In part, this reflects the

history of oyster-related enhancement activities directed

towards reef construction to support commercial harvest,

highlighting the unique peril of human consumptive dis-

turbance for shellfishes. Recent socio-economic analyses

have argued, however, that unharvested restored reefs can

return five times greater value in goods and services than

reefs exploited solely for shellfish production (Grabowski

& Peterson 2007).

Subsequently, efforts to enact long-term oyster restora-

tion have remained chiefly focused on subtidal reefs, with

some notable successes in merging large-scale restoration

with basic hypothesis testing and model development (e.g.

North et al. 2010). Experimental restoration in Pamlico

Sound, North Carolina (NC, 1990s), and Chesapeake

Bay, Virginia (2000s), explored how oyster-reef relief

interacted with the physicochemical environment. These

studies showed that taller reefs are less susceptible to the

catastrophic impacts of bottom-water hypoxia (Lenihan &

Peterson 1998; Lenihan 1999), thus improving physiologi-

cal condition and also minimizing the deleterious effects

of disease and sedimentation (Schulte, Burke & Lipcius

2009). Consequently, managing reef relief has become a

priority for both exploited and restored oyster reefs.

Intertidal oysters have been targeted less frequently for

restoration, despite the fact that those reefs once com-

prised a significant portion of overall reef habitat in euha-

line waters (Winslow 1886). In NC, some restored

intertidal reefs have exceeded milestone thresholds (e.g.

live densities), while other reefs sited just 10s–100s of

metres away from these successful projects have failed

completely (e.g. NC Division of Marine Fisheries enhance-

ment sites 95-965 and 96-049, J. Fodrie, personal observa-

tions). These contrasting outcomes highlight gaps in our

understanding or application of the biotic and abiotic

mechanisms governing reef ecology, requiring improved

models to more effectively intervene in nature and achieve

restoration goals.

Multiple lines of evidence suggest that the trajectory of

restored oyster reefs could be predicted by considering

paradigms best known from rocky shores and saltmars-

hes. Reports dating from the 1800s have posited that in

some environments, oysters can be restricted to the inter-

tidal as a refuge from predation, biofouling, bioerosion,

disease or sedimentation (reviewed in Bahr & Lanier

1981) – analogous to rocky shores. Furthermore, oyster

growth on aquaculture racks can be higher intertidally vs.

subtidally, perhaps due to decreased biofouling (i.e.

smothering; Bishop & Peterson 2006), and akin to Sparti-

na production across elevations in saltmarsh ecosystems.

Still, to our knowledge, no explicit tests exist to assess

how aerial exposure regulates restored reef performance

and thus guide future restoration activities. This led us to

restore reefs across the shallow-subtidal-to-mid-intertidal

exposure gradient and evaluate: (i) do gradients or thresh-

olds emerge in oyster densities and reef succession across

depths (over cm scales in the vertical); (ii) for reefs that

appear different, over what time-scales do patterns emerge

© 2014 The Authors. Journal of Applied Ecology © 2014 British Ecological Society, Journal of Applied Ecology, 51, 1314–1325

Littoral ecology guides biogenic reef restoration 1315

and what mechanisms potentially drive these gradients;

and (iii) does reef size (footprint) mitigate the effects of

depth by decreasing the relative amount of edge habitat

where factors such as predation or sedimentation may be

strongest?

Materials and methods

STUDY SITE

To test depth (aerial exposure) and size effects on the succession

of restored oyster reefs via natural processes of settlement,

growth and interspecific interactions, we constructed 32 oyster

shell piles in four zones encircling Middle Marsh, NC, a relic

flood-tide delta located within the Rachel Carson National Estua-

rine Research Reserve (Fig. 1). Reefs were constructed on broad

sandflats surrounding this 1!5-km2 saltmarsh complex, while nat-

ural reefs exist in Middle Marsh on similar tidal flats or fringing

shorelines (Grabowski et al. 2005). In the Reserve, the tidal range

extends from "0!75 to 0!73 m relative to the North American

Vertical Datum of 1988 [NAVD88; mean sea level (50% expo-

sure): "0!03 m NAVD88] based on long-term water-level data

collected by on-site pressure gauges (Onset HOBO U20 loggers).

EXPERIMENTAL DESIGN AND REEF CONSTRUCTION

Within each zone surrounding Middle Marsh, we identified eight

locations for creating reefs. Four reefs within each zone were con-

structed using 60 bushels of oyster shell per reef with initial dimen-

sions of 3*5*0!15 m (W*L*H) while the remaining four reefs were

built using 300 bushels of oyster shell per reef with initial dimen-

sions of 8*10*0!15 m. Inside each zone, we sited pairs of small

and large reefs at four different depths relative to NAVD88: "0!5,"0!6, "0!75 and "0!9 m (at their base). Reefs at the shallowest

two depths were exposed during every low tide, while reefs con-

structed at "0!9 m NAVD88 were nearly always inundated. Reefs

at "0!75 m NAVD88 were built at the level of approximate

monthly mean low water, and aerial exposure depended on tidal

magnitude (spring vs. neap tides) and meteorological conditions

(NE winds increase water depth at Middle Marsh relative to astro-

nomical predictions, while SW winds reduce water levels). Expo-

sure periods for reefs at each depth (from base to crest) were

calculated from water-level data as: "0!5 m = 5!0–17!9%exposure; "0!6 m = 1!5–8!3%; "0!75 m = 0!1–1!5%; and

"0!9 m ≤ 0!1%.

During March 2011, we identified reef construction sites that

were precisely at those four target depths. Each study zone was

mapped using a Trimble 5800 Real Time Kinematic Global Posi-

tion System (RTK-GPS; <1!5-cm vertical precision). The resulting

bathymetric maps were based on regular transects within each

zone and guided reef placement (20- to 30-m spacing across the

entire sandflat, with targeted 2-m spacing inside focal areas). Ulti-

mately, the R2 between planned and actual reef depths was 0!97,and the vertical offset for reefs relative to our intended design

averaged <3 cm. All reefs were spaced >30 m apart to minimize

the potential for one restored reef to affect nearby experimental

reefs. Oyster shells (‘cultch’; >7!5 cm long) were bought from an

oyster-shucking company in the fall of 2010 and allowed to ‘cure’

for several months to preclude faunal/floral translocation. Reefs

were constructed in May 2011 using the general methods

described by Grabowski et al. (2005), which involved dumping

shells from a barge, and then shaping cultch piles into uniform

rectangular dimensions using hand tools and a modified dredge.

FIELD SAMPLING

Reefs were sampled three times within the first year following con-

struction to explore patterns of oyster settlement, density, survival

and sizes, as well as reef-associated faunal densities across our

experimental exposure gradient. Sampling occurred in July and

September 2011 and May 2012 based on the expected early summer

Fig. 1. Restored oyster reefs surroundingMiddle Marsh, NC. The upper-left imageshows the regional setting and four repli-cate zones where reefs were built. Withineach zone, bathymetric maps were created(images at right and bottom) to guide thesiting of small and large cultch shell pilesat one of four depths: "0!5, "0!6, "0!75and "0!9 m NAVD88.


1316 F. J. Fodrie et al.

peak in oyster settlement and the seasons most suitable for measur-

ing oyster growth and survival (Ortega & Sutherland 1992).

During our first sampling effort, we recorded oyster (spat) set-

tlement, exposed oyster shell cover, algal cover and counts of

other fouling organisms such as barnacles Semibalanus balanoides

and bryozoans Bugula neritina. On each reef, shell cover and

algal cover were recorded within two randomly selected 0!25-m2

quadrats using a standard grid-intersect approach. Because cultch

shells were still quite loose and small spat were nearly impossible

to count on the rough exterior of each shell, we counted spat and

other fouling organisms inside the quadrat only from the inward-

facing half of cultch oyster shells. Nearly, all spat were <0!25 cm

in length (all were <1!0 cm); therefore, no length measurements

were made during the first sampling period.

During our September sampling, we quantified oyster densities,

exposed cultch shell cover, algal cover and densities of other reef-

associated organisms. We also recorded evidence of oyster mor-

tality, sized individual oysters and examined the basic flow regime

at each reef. As before, we located two random 0!25-m2 quadrats

on each reef and quantified shell cover and algal cover. By

September, oysters had cemented together considerably making it

impossible to remove individual cultch shells for density counts.

Therefore, we excavated the top 10 cm of material within each

0!25-m2 quadrat. Samples were sieved through a 1-mm mesh, and

every live oyster was counted and measured to the nearest mm

from the hinge to the leading growth margin. Visible spat scars

(dead oysters) were also counted but not measured. Oyster mor-

tality was calculated as the percentage of scars relative to the

sum total of live oysters and scars. Gastropods, xanthid crabs

(mud crabs), mytilid bivalves (mussels) and soft-bodied infauna

were the major taxa or functional groups we encountered in sam-

ples. Therefore, those organisms were separated, identified and

enumerated. Due to inclement weather in September 2011, we did

not sample the restored reefs in zone three.

During September 2011, we used dissolution blocks to examine

potential gradients in flow regimes among reefs that could be cor-

related with differences in sedimentation or oyster fitness. We

constructed 64 gypsum cylinders 10*1 cm (D*H) and covered the

side and one end of each cylinder with two layers of polyurethane

to ensure that an equal surface area would be subject to dissolu-

tion at all times. Each cylinder was weighed (g) and then adhered

with silicone to a brick (20*20*4 cm) that could be placed flush

on top of reefs. Two gypsum cylinders were randomly deployed

on each reef and recovered 2!5 tidal cycles later. Following retrie-

val, gypsum cylinders were dried for 24 h and reweighed.

In May 2012, we quantified oyster densities, exposed cultch

shell cover, algal cover and presence of other reef-associated

organisms following our September 2011 protocols. We also sized

all live oysters in each quadrat, but were not able to record evi-

dence of oyster mortality as scars representing the 2011 cohort

were indistinguishable due to oyster overgrowth and continued

biofouling. Additionally, we walked the perimeter of each reef

and counted the number of active Menippe spp. (stone crab;

4–10 cm carapace length) burrows, which were present along the

reef–sediment margin and easily identifiable by the shell frag-

ments at each burrow entrance.

STATIST ICAL ANALYSES

Settlement patterns during July 2011 were analysed by regressions

in which spat counts (# 0!25 m"2) were plotted against

sedimentation (% shell cover), barnacle densities (# 0!25 m"2)

and bryozoan densities (# 0!25 m"2). For each sampling period,

we also compared oyster densities among reefs at different depths

and of different sizes via two-way analyses of variance (ANOVAS).

Because of the different methods employed between the July and

later collections, separate ANOVAs and interpretations were

required for each sampling period (i.e. a repeated-measures

design was precluded).

We used two-way ANOVAs to test for effects of reef depth and

size on shell cover, algal cover, stone-crab burrow density (stan-

dardized by overall reef area and length of reef perimeter), gas-

tropod density, mud-crab density, mussel density and soft-bodied

infaunal density. Analyses of reef-associated fauna were run only

for the May 2012 collections as very few individuals were col-

lected from earlier surveys (less than five individuals per sample).

We used Kruskal–Wallis tests to determine the independent

effects of reef depth and size on oyster lengths (a potential proxy

for individual growth) during September 2011 and May 2012. We

also tested for differences in oyster mortality or gypsum dissolu-

tion among reefs measured in September 2011 using two-way

ANOVAs with reef depth and size as factors. Additionally, we

examined the relationship between oyster mortality and sedimen-

tation (shell cover) at the quadrat level using linear regression.

Data were tested for normality and homogeneity of variances

(Fmax test) for each main effect. Shell cover, oyster mortality,

gypsum dissolution and gastropod densities were analysed with-

out transformations. Heterogeneous data were rectified with log

(x + 1) [densities of oysters, burrows (per reef area and perime-

ter), mud crabs, mussels and soft-bodied infauna] or arcsine

(algal cover) transformations to stabilize variances. Transforma-

tions failed to produce homogeneous variances for oyster size

data, and therefore, main effects were tested using nonparametric

statistics. In cases with significant main effects (a < 0!05), we used

Fisher’s (parametric) or Mann–Whitney U (nonparametric) tests

for post hoc comparisons between treatments. Given the scale of

our restored reefs, we ran our analyses with each 0!25-m2 quadrat

as a replicate.

Results

Oyster densities were significantly affected by reef depth

(aerial exposure), but with a reversal over time in density

gradients across depths (see Table S1, Supporting Infor-

mation; Fig. 2). In July 2011, oyster-spat counts on the

shallowest reefs were only 20–25% of those recorded on

the deepest two reef sets (c. 3000 oysters m"2 of spread

cultch shell), with ‘intermediate’ spat settlement on the

"0!6-m reefs (P = 0!007). At the quadrat level, settlement

did not scale with sedimentation (Fig. S1A; R2 = 0!13) orbarnacle density (Fig. S1B; R2 = 0!11). Spat settlement

scaled positively with bryozoan density, as bryozoans

were limited to the deeper reefs where spat counts were

also highest (Fig. S1B; R2 = 0!39). In September 2011,

however, oyster counts on the "0!9-m reefs were signifi-

cantly less than on the three shallower treatments

(P < 0!001). Indeed, a complete reversal was observed, as

we recorded c. 2500 oysters m"2 on the deepest reefs, and

roughly 10 000 oysters m"2 on all other reefs. In May

2012, we recorded a clear pattern of high-to-low oyster

densities from the shallowest-to-deepest reefs (P = 0!002).



Oyster densities approached 2200 oysters m"2 on reefs at

"0!5 m, but declined across reefs at "0!6 (c. 1250 oys-

ters m"2), "0!75 (c. 800 oysters m"2) and "0!9 m

(c. 400 oysters m"2) (Fig. 3). We did not detect any

significant main or interactive effects of reef size on

oyster densities throughout this study (Table S1; all

P-values > 0!129).Similarly, shell cover was affected by reef depth (all

P-values < 0!009), but not reef size (all P-values > 0!106),during each sampling period (Table S2; Fig. 4a,c,e). Dur-

ing July and September 2011, sediment cover was signifi-

cantly greater on the deepest reefs. Shell cover averaged

50–65% on the reefs at "0!9 m, while shell cover was

greater (75–95%) on shallower reefs. By May 2012, the

two deeper reef sets were characterized by c. 50% shell

cover, while the shallower two reef sets were defined by

significantly greater shell cover (70–75%). Several algal

species were observed on our restored reefs during 2011–2012 (from most-to-least abundant): Ulvaintestinalis,

Ectocarpus spp., Ceramium spp., Chondria spp., Hypnea

spp., Lomentaria spp., Agardhiella subulata, Gracilaria

verrucosa, Champia spp. and Dictyota menstrualis. The

effects of reef depth and size on algal cover varied through

time (Table S2), but in general, algal cover decreased as

aerial exposure among reefs increased (Fig. 4b,d,f). In July

2011, the "0!9-m reefs were characterized by significantly

higher algal cover (c. 15%) than the shallower three reef

sets (P = 0!001). Two months later, all reefs had little

(<10%) algal cover (P = 0!180). In May 2012, we recorded

the only reef depth*size interaction in our entire data set

(P = 0!006). On small reefs, algal cover increased linearly

with depth from <5% to nearly 40%, while on large reefs

we observed peak cover (50%) at "0!75 m.

By September 2011, we recorded significant differences

in oyster lengths among depths (Fig. 5a; P < 0!001). Thelargest oysters (c. 18 mm) were on the reefs at "0!6 m,

followed by individuals on the "0!5-m reefs (c. 15 mm).

Oysters at "0!75 and "0!9 m were c. 33% smaller

(10–12 mm). In May 2012, all four depth treatments

could be distinguished by mean oyster size (P < 0!001).Again, the largest oysters were on the reefs at "0!6 m

(c. 45 mm), while the reefs at "0!5 m were also defined

by relatively large oysters (c. 35 mm). The oysters at

"0!75 and "0!9 m were roughly 18 and 22 mm, respec-

tively, or 37–60% smaller than the oysters on the shallow-

est two treatments. Depth also had a significant effect on

oyster mortality in September 2011 (P = 0!003), althoughpatterns relative to depth were not straightforward

(Table S3, Fig. 5b). The "0!5- and "0!75-m reefs were

characterized by lower mortality rates (25–35%) relative

to the reefs at "0!6 and "0!9 m (40–50%). We noted

that mortality rates were variable, although not statisti-

cally significant, among large and small reefs (depth*size

interaction P = 0!069), and that for large reefs there was

an obvious trend of increasing mortality with increasing

reef depth. We recorded a modest correlation between

increasing shell cover and decreasing mortality rates

within individual quadrats (Fig. 5c; R2 = 0!34). We did

not detect any statistically significant effects of reef depth

or size on the rate of gypsum dissolution (Table S3).

As of May 2012, reef depth had significant effects on

the density of three out of five reef-associated invertebrate

0 1000 2000 3000 4000 5000 6000

–0·50

–0·60

–0·75

–0·90

Oysters per m2 of spread shell (ђ + 1 SE)

Oysters per m2 (ђ + 1 SE)

Oysters per m2 (ђ + 1 SE)

Large reefs

Small reefs

July 2011

a

Dep

th a

t ree

f bas

e (m

; NAV

D88

)

c

b

b,c

0 5000 10 000 15 000

–0·50

–0·60

–0·75

–0·90

Sept 2011

a

Dep

th a

t ree

f bas

e (m

; NAV

D88

)

a

a

b

0 1000 2000 3000 4000

–0·50

–0·60

–0·75

–0·90

May 2012

a

a,b

b,c

c

Dep

th a

t ree

f bas

e (m

; NAV

D88

)

(a)

(b)

(c)

Fig. 2. Oyster densities in (a) July 2011, (b) September 2011 and(c) May 2012. Direct comparisons of oyster densities across timewere not possible given the different sampling techniques used tocount spat (July 2011) and larger oysters (September 2011 andMay 2012). Bars represent the mean of six (September 2011) oreight (July 2011 and May 2012) replicate quadrats sampledamong reefs (+1 SE). Significant differences (a < 0!05) betweendepths based on post hoc analyses are represented by different let-ters at the right of the bars.



taxa, although the direction of change across our vertical

exposure gradient was taxon-specific (Table S4; Fig. 6).

Stone-crab burrow densities per-unit-area-reef were signifi-

cantly (independently) affected by reef depth (P < 0!001)and size (P = 0!002), while burrow counts per-length-reef-

perimeter were statistically different only among depths

(P < 0!001). Burrow densities were highest on the reefs

at "0!75 m, and 75% (per-area-reef) to 80% (per linear

edge) lower on the reefs at "0!5 and "0!6 m (intermediate

burrow densities on the deepest reefs). Several gastro-

pod species that are known oyster predators (Bahr &

Lanier 1981) were collected from our restored reefs,

including (from most-to-least abundant): Fasciolaria

lilium, Urosalpinx cinerea, Eupleura caudate, Busycon

carica, Busycotypus canaliculatus and Busycon sinistrum.

Collectively, we did not detect any statistically significant

effects of reef depth (P = 0!207) or size (P = 0!501) on the

density of predatory gastropods. We did, however, note a

statistically non-significant trend of increasing mean gas-

tropod density with increasing depth: the deepest reefs

(>30 gastropods m"2) were, on average, occupied by three

times more gastropods than the shallowest reefs (<10 gas-

tropods m"2). Densities of mud crabs, dominated by pur-

ported oyster consumers such as Eurypanopeus depressus

and Panopeus herbstii (Grabowski 2004), were signifi-

cantly affected by reef depth (P = 0!009). As many as

–0·9

m–0

·75

m–0

·60

m–0

·50

m

Large reef (8×10 m) Small reef (5×3 m) Ground-Level view

1 m

Fig. 3. Overhead (all treatments) and ground-level (large reefs only) images of restored reefs in Zone 2 across intertidal ("0!5, "0!6 and"0!75 m NAVD88) and shallow subtidal ("0!9 m NAVD88) settings nearly 12 months post-construction.



150 crabs m"2 were collected from the "0!5-m reefs, while

the density of crabs decreased with depth by fivefold to

c. 30 crabs m"2 on the "0!9-m reefs. Similarly, the den-

sity of mytilid mussels (mostly Geukensia demissa and

Brachidontes exustus) decreased significantly with depth

(P < 0!001), ranging from c. 1500 mussels m"2 on the

"0!5-m reefs to <30 mussels m"2 on the reefs constructed

at "0!9 m (a 50-fold change). Nereid worms dominated

the soft-bodied infauna, and we recorded no significant

effects of reef depth (P = 0!639) or size (P = 0!093) on

this reef-associated group (uniformly 175–200 worms m"2

among treatments).

Discussion

Natural and restored shellfish reefs are increasingly

valued for their potential to deliver a suite of ecosystem

goods and services related to finfish production

(Grabowski & Peterson 2007), water filtration (North

et al. 2010) and shoreline protection (Meyer, Townsend &

Thayer 1997). Many oyster-reef restoration projects, how-

ever, have met with limited success owing, in part, to our

incomplete understanding or application of the regulatory

mechanisms that govern reef ecology across estuarine and

regional scales (Coen & Luckenbach 2000; Powers et al.

0 10 20 30 40 50 60 70 80 90 100

–0·50

–0·60

–0·75

–0·90

Large reefs

Small reefs

July 2011

b

a

a

a

0 10 20 30 40 50 60 70 80 90 100

–0·50

–0·60

–0·75

–0·90

July 2011

a

a

a

b

Dep

th a

t ree

f bas

e (m

; NA

VD

88)

0 10 20 30 40 50 60 70 80 90 100

–0·50

–0·60

–0·75

–0·90

Sept 2011

a

b

a

a

Dep

th a

t ree

f bas

e (m

; NA

VD

88)

0 10 20 30 40 50 60 70 80 90 100

–0·50

–0·60

–0·75

–0·90

% algal cover (ђ + 1 SE)

May 2012

0 10 20 30 40 50 60 70 80 90 100

–0·50

–0·60

–0·75

–0·90

% shell cover (ђ + 1 SE)

a

a

b

b

Dep

th a

t ree

f bas

e (m

; NA

VD

88)

May 2012

0 10 20 30 40 50 60 70 80 90 100

–0·50

–0·60

–0·75

–0·90

Sept 2011

(a) (b)

(c) (d)

(e) (f)

Fig. 4. Percentage cover of (a, c, e) shell and (b, d, f) algae in (a, b) July 2011, (c, d) September 2011 and (e, f) May 2012. Bars representthe mean of six (September 2011) or eight (July 2011 and May 2012) replicate quadrats sampled among reefs (+1 SE) based on a grid-intercept approach. Post hoc comparisons between depth treatments were made whenever that main effect was significant, with signifi-cant differences (a < 0!05) between depths represented by different letters at the right of bars (but not possible for the depth*size interac-tion in panel f).



2009). In littoral environments, we considered whether

food web and productivity paradigms that operate along

pronounced vertical gradients – well studied previously in

rocky shore and saltmarsh ecosystems – could be applied

to optimize the design of restoration activities involving

shellfish reefs. We showed that: (i) gradients in spat settle-

ment were decoupled from subsequent vertical patterns in

oyster densities. We hypothesize this was due to biotic

controls (predation and competition), which increased in

intensity with depth (i.e. decreased aerial exposure;

Fig. 7). These forces required <1 year to reverse initial

settlement patterns and generate strong gradients in reef

development; (ii) for 15-cm-tall reefs, and regardless of

reef size, a threshold in reef performance existed at

"0!6 m NAVD88 (equivalent to c. 5% daily exposure).

Below this depth, restored reefs failed; and (iii) just above

this "0!6-m threshold, a vertical ‘hot spot’ of oyster bio-

mass accumulation occurred (Fig. 7).

OYSTER REEFS AS ROCKY INTERTIDAL AND

SALTMARSH ANALOGUES

We exploited this restoration effort as a tool for critically

evaluating whether oyster-reef communities are direct

analogues to rocky and marshy littoral systems, and if

transferrable lessons can guide restoration. While no gen-

eral consensus regarding vertical patterns of oyster settle-

ment exists (e.g. Chestnut & Fahy 1953; McNulty 1953;

Bartol & Mann 1997), we observed strong vertical gradi-

ents in settlement among our restored reefs. Settlement

patterns, however, were decoupled from the densities of

larger oysters <1 year later. These findings mirror the dis-

tribution of larval and adult stellate barnacles on rocky

shores (Connell 1961; Wethey 1984; Fig. 7).

The primacy of competition for space in rocky and salt-

marsh littoral systems is well-established (Connell 1961;

Bertness & Silliman 2014). In restored oyster reefs, we

found modest evidence that competitive exclusion contrib-

uted to vertical gradients in oyster densities (Fig. 7). Semi-

balanus barnacles began settling on restored reefs in late

May and early June 2011 (prior to oyster settlement) and

were most abundant on shallow reefs, but settled oysters

were ultimately able to become the numerically dominant

fouling organism at all depths. Conversely, oysters them-

selves face the threat of smothering from algal mats and

bryozoans, which were limited to deeper restored reefs

due to the physiological costs of desiccation. Bishop and

Peterson (2006) found that despite significant aerial expo-

sure, intertidal Crassostrea ariakensis grew more rapidly

than subtidal conspecifics and hypothesized that decreased

biofouling on periodically exposed oysters explained this

paradoxical result. Likewise, we found that the sizes of

surviving oysters on the deeper two reef sets were only

half as large as individuals on the shallower two sets, sug-

gesting potential for restricted growth in subtidal environ-

ments (but see competing hypotheses below).

Top-down effects also profoundly affect rocky intertidal

food webs and can interact with abiotic stressors to impact

saltmarsh communities (Paine 1966; Gittman & Keller

2013). Our observational data lead us to hypothesize that

predation by crustaceans and gastropods likely played a

significant role in generating differences among depth

treatments (Johnson & Smee 2014; Fig. 7). Our inspec-

0 10 20 30 40 50

–0·50

–0·60

–0·75

–0·90

Oyster shell length (mm; µ + 1 SE)

May-12Sep-11a

b

Cc

d

(a)

A

C

B

Dept

h at

reef

bas

e (m

; NAV

D88)

0 10 20 30 40 50 60 70 80

–0·50

–0·60

–0·75

–0·90

Oyster mortality (% visible scars; µ + 1 SE)

Dept

h at

reef

bas

e (m

; NAV

D88)

Large reefsSmall reefs

(b)

a

b

b

a

0102030405060708090

100

0 25 50 75 100

Oyst

er m

orta

lity

(% v

isibl

e sc

ars)

% shell cover in sampling quadrat

R2 = 0·34P < 0·001

(c)

Fig. 5. (a) Oyster lengths in September 2011 and May 2012. Barsrepresent the mean lengths of all oysters found within six (Sep-tember 2011) or eight (May 2012) replicate quadrats (+1 SE).(b) Oyster mortality across depths, calculated as the percentageof scars relative to total live and dead oysters, observed in Sep-tember 2011. Bars represent the mean of six replicate mortalityestimates from quadrats (+1 SE). (c) Relationship between oystermortality and shell cover within quadrats. In (a), upper- andlower-case letters for September 2011 and May 2012 compari-sons, respectively, and (b), significant differences (a < 0!05)between depths based on post hoc analyses are represented by dif-ferent letters at the right of the bars.



tions of oyster scars revealed that the majority of dead

individuals were characterized by missing or broken right

valves, indicative of mechanical injury rather than abiotic

stress. Surveys of stone-crab burrows in May 2012 showed

that this predator was predominantly occupying the reefs

at "0!75 and "0!9 m. Although not statistically signifi-

cant, faunal sampling also suggested that predatory gas-

tropods were more abundant on our deeper reefs. These

gastropods forage more effectively when submerged and

were regularly observed feeding on oysters within deeper

reefs throughout the fall of 2011. Conversely, mud-crab

densities in May 2012 were greatest within intertidal reefs

where oyster densities were also highest. While mud crabs

may prey upon newly settled oysters (Grabowski 2004),

they were not among the early, successful colonizers of

our experimental reefs. Therefore, we suspect mud crabs

did not exert top-down regulation on the 2011 oyster

cohort. Rather, mud crabs sampled in May 2012 were

0 500 1000 1500 2000 2500 3000

–0·50

–0·60

–0·75

–0·90

MyƟlid mussels per m2 ( ђ + 1 SE)

bDep

th a

t ree

f bas

e (m

; NA

VD88

)

a

a

b

0 50 100 150 200 250

–0·50

–0·60

–0·75

–0·90

Mud Crabs per m2 (ђ + 1 SE)

a,b

a

a,c

c

0 10 20 30 40 50 60

–0·50

–0·60

–0·75

–0·90

Predatory Gastropods per m2 (ђ + 1 SE)

Dep

th a

t ree

f bas

e (m

; NA

VD88

)

0 100 200 300 400 500 600 700

–0·50

–0·60

–0·75

–0·90

SoŌ-bodied infauna per m2 (ђ + 1 SE)

0 0·1 0·2 0·3 0·4

–0·50

–0·60

–0·75

–0·90

Stone crab burrow density per m2 (ђ + 1 SE)

Large reefs

Small reefs

Dep

th a

t ree

f bas

e (m

; NA

VD

88)

a

a

b

c

0 0·05 0·1 0·15 0·2 0·25 0·3

–0·50

–0·60

–0·75

–0·90

Stone crab burrow density per m edge (ђ + 1 SE)

a

a

b

c

(e)

(d)(c)

(f)

(a) (b)

Fig. 6. Density of (a) stone-crab burrows per reef area, (b) stone-crab burrows per linear edge of reef, (c) gastropods, (d) xanthid crabs,(e) mytilid mussels and (f) soft-bodied infauna in May 2012. Bars represent the means of replicate surveys (a, b; n = 4) or quadrats(0!25 m"2; c–f; n = 8) (+1 SE). Significant post hoc comparisons between depths (a < 0!05) are represented by different letters at the rightof bars.



most likely relying on the structural refuge provided by

developing reefs, similarly to mytilid mussels, while con-

suming algal and detrital material (Bahr & Lanier 1981).

Other factors such as bioerosion, disease and sedimenta-

tion could contribute to vertical gradients in shellfish fit-

ness, although they are not considered strong regulatory

mechanisms in rocky intertidal systems (Fig. 7). Cliona spp.

(boring sponge), Perkinsus marinus (Dermo) or Haplospori-

dium nelsoni (MSX) can result in >90% mortality on indi-

vidual oyster reefs within euhaline environments (Hopkins

1962; Newell 1988). Typically, however, these pests and

protozoan parasites require gestations of 2–3 years before

they impact newly infected oysters. Furthermore, we saw

no signs of Cliona in any living or cultch oyster shells, and

regional disease rates are low (Powers et al. 2009).

Smothering via sedimentation has affected restored reef

performance in previous reports. In Delaware Bay,

low-relief, restored, intertidal reefs were completely buried

by shifting sediment banks, while high-relief reefs sup-

ported oyster-reef communities (Taylor & Bushek 2008).

We found no relationship between sedimentation and spat

settlement, and throughout our study (especially in May

2012), sediment cover did not scale with differences in oys-

ter densities, mortality rates or gypsum dissolution. Given

the absence of gradients in bulk flow or links between sedi-

ment cover and oyster fitness, we suspect that buried shell

was a symptom of reef failure in our system, but not a

mechanistic cause. Rather, intertidal reefs in our study

were characterized by a thin (5- to 10-cm) veneer of verti-

cally extending live oysters in the taphonomically active

zone that simply did not exist on deeper reefs.

We do caution that our experiment was not designed to

separate the effects of reef depth (at their base) vs. reef

relief, and the potential for an interaction between these

factors remains unresolved. For example, while a

15-cm-tall reef constructed at "0!75 m NAVD88 failed in

our study, we do not suggest that a 30-cm-tall reef con-

structed at that same depth would also fail. Perhaps oysters

would persist on the portion of that reef (or any reef) above

the "0!6-m NAVD88 threshold, and we do anticipate that

predation rates could be reduced on high-relief reefs regard-

less of depth (sensu Lenihan 1999). Furthermore, dislodged

oyster clumps from the portion of the reef above "0!6 m

NAVD88 could subsidize deeper portions of the reef.

Revisiting our oyster size data highlights potential par-

allels between ‘production’ in oyster reefs and saltmarsh

ecosystems. At least three factors could have contributed

to the differences we observed in mean oyster sizes among

depths: first, growth rates of oysters may have varied

across aerial exposures (sensu Bishop & Peterson 2006);

secondly, the tight clustering of surviving oysters in the

intertidal could have resulted in elongated growth mor-

phologies on those reefs (sensu Bahr & Lanier 1981); and

thirdly, predators may have disproportionately consumed

larger oysters on the subtidal reefs (and to a lesser degree,

the reefs at "0!5 m NAVD88). While our experiments

could not distinguish these mechanisms, we can consider

the ultimate consequences and parallels of oyster biomass

accumulation (density*size) across depths. Morris et al.

(2002) found that Spartina production in South Carolina

marshes was greatest at roughly "0!6 m below mean high

tide (in South Carolina: 0!15 m NAVD88, or c. 55%

exposure during the tidal cycle). At that elevation, peri-

odic exposure alleviated abiotic stress for plants while reg-

ular inundation allowed sufficient nutrient and sediment

delivery. Previous experimental work by Bishop and

Larvae Larvae

Adults AdultsDesiccation Desiccation, food & larval limitation

Chthamalus stellatus Crassostrea virginica

Distribution Distribution

Growth (biomass accumulation)

A B A B C

A: Predation byThais

B: Competition withBalanus

A: Predation by Menippe, Busycon, etc.B: Competition/smothering by fouling

organisms (macroalgae, etc.)C: Sedimentation, bioerosion and disease

?

SHW

MHW

MW

MLW

SLW

Mar

ine

B

rack

ish(N

ot a

pplie

d to

des

iccat

ion

or fo

od fl

ux)

Fig. 7. Processes that generate zonation for a rocky intertidal barnacle, Chthamalus stellatus (taken from Connell 1961), also regulatethe trajectory of restored oyster reefs in Middle Marsh, NC.



Peterson (2006) supports the hypothesis that differential

growth of oyster across depths could have contributed to

the parabolic pattern of oyster sizes we recorded, reflect-

ing gradients in biofouling, predation threat or abiotic

stress (desiccation) (Fig. 7: analogous to saltmarsh pro-

duction). This growth difference (including potential gra-

dients in growth morphologies or size-specific predation),

combined with the observed differences in oyster densities

among depths, resulted in greatest accumulation of new

oyster biomass on intertidal reefs in our study. Presuming

that oyster biomass accumulation scales positively with

vertical reef growth, unimodal patterns of biogenic habitat

accretion across depths are likely shared among oyster

reefs and saltmarshes despite different forcing mechanisms

(Redfield 1972).

GUIDING REEF RESTORATION

The literature is clear that early researchers studying oys-

ter reefs had a general appreciation for the paradigms of

community organization well known from other zonated

littoral systems. For instance, competition among barna-

cles, mussels and oysters on vertical pilings was studied

throughout the 1960s and 1970s to explain zonation in

fouling communities across depths and wave-energy gradi-

ents (e.g. Ortega 1981). Even earlier, high predation rates

(e.g. oyster drills; Galtsoff 1964) and bioerosion (Hopkins

1962) were suspected to limit oysters in many euhaline

sites to intertidal refugia (Winslow 1886).

Still, several factors have led to the apparent omission of

intertidal paradigms as guides for improving biogenic reef

restoration. The most insidious challenge is that natural

reefs have been dramatically reduced in extent and biomass

(Zu Ermgassen et al. 2012). Remaining reefs face interact-

ing stressors such as harvest pressure, diminished water

quality (hypoxia, sedimentation, toxins), climate change

(heat stress, saltwater intrusion) and localized disturbance

(boat wakes) (Beck et al. 2011). Thus, the natural history of

shellfish reefs is no longer readily discerned – a problem

exacerbated by any under-appreciation for early research in

these systems.

Furthermore, the threshold between restored reef

success or failure can span a vertical distance of only

10–15 cm. This elevation change can only be discerned

using GPS-based systems with high vertical resolution;

therefore, improved shellfish restoration could arise sim-

ply by marrying food web and productivity paradigms

with proper surveying technology.

To support conservation efforts, bioeconomic models

are emerging that relate oyster biomass and reef connec-

tivity with the provision of key ecosystem services (North

et al. 2010; Grabowski et al. 2012). In turn, improved

ecological models that quantify gradients or thresholds in

oyster fitness in response to settlement, predation, compe-

tition, bioerosion, disease, hypoxia/anoxia and sedimenta-

tion should underpin these analyses to guide enhanced

restoration practices. In littoral systems, our results com-

pliment previous findings relating reef performance to ver-

tical relief (Taylor & Bushek 2008) or landscape setting

(Grabowski et al. 2005), and prescribe a vertical ‘hot spot’

for restoration to maximize biogenic reef fitness and pro-

ductivity that will in turn support the delivery of a suite

of ecosystem services.

Acknowledgements

We thank A. Tyler and the NC Division of Marine Fisheries for assis-tance in constructing reefs. We are also indebted to the technicians andundergraduates who helped with reef shaping or sampling, especiallyR. Bouchillon and K. Ellis. D. Kimbro and two anonymous reviewersprovided comments that substantially improved this manuscript. Thisresearch was supported by funding from the Albemarle-Pamlico NationalEstuary Partnership, NC Sea Grant, NC Marine Resources Fund and theNational Science Foundation (OCE-1155628). We declare no competinginterests.

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Received 25 August 2013; accepted 23 April 2014Handling Editor: Shelley Arnott

Supporting Information

Additional Supporting Information may be found in the online version

of this article.

Table S1. ANOVA table for reef depth and size effects on oyster

densities.

Table S2. ANOVA table for reef depth and size effects on shell and

macroalgal cover.

Table S3. ANOVA table for reef depth and size effects on oyster

mortality and gypsum dissolution.

Table S4. ANOVA table for reef depth and size effects on reef-associated fauna.

Fig. S1. Spat settlement vs. sedimentation, barnacle density and

bryozoan density.



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