Next Article in Journal
Computational Analysis of Stiffness Reduction Effects on the Dynamic Behaviour of Floating Offshore Wind Turbine Blades
Next Article in Special Issue
Fan Mussel (Pinna nobilis L.) Spat Collection, Monitoring of Early Growth and Conservation Implications by Deploying Conventional Aquaculture Methodology
Previous Article in Journal
UUV-Assisted Icebreaking Application in Polar Environments Using GA-SPSO
Previous Article in Special Issue
Spatial and Temporal Variability in Oyster Settlement on Intertidal Reefs Support Site-Specific Assessments for Restoration Practices
 
 
Search for Articles:
Title / Keyword
Author / Affiliation / Email
Journal
Article Type
 
 
Advanced Search
 
Section
Special Issue
Volume
Issue
Number
Page
 
Journals
JMSE
Volume 12
Issue 10
10.3390/jmse12101847
jmse-logo
Submit to this Journal Review for this Journal Propose a Special Issue
► Article Menu

Article Menu

share Share announcement Help format_quote Cite question_answer Discuss in SciProfiles thumb_up
...
Endorse textsms
...
Comment
first_page
settings
Order Article Reprints
Font Type:
Arial Georgia Verdana
Font Size:
Aa Aa Aa
Line Spacing:
Column Width:
Background:
Article

Challenges to Seagrass Restoration in the Indian River Lagoon, Florida

by
Vivienne A. Main
1,2,*,
Morgan K. Gilligan
1,3,
Sarah M. Cole
1,4,
Todd Z. Osborne
5,
Ashley R. Smyth
6 and
Loraé T. Simpson
1,7
1
Florida Oceanographic Society, 890 NE Ocean Blvd, Stuart, FL 34996, USA
2
Centre Global, Centre College, Danville, KY 40422, USA
3
Pacific Northwest National Laboratory, 902 Battelle Blvd, Richland, WA 99354, USA
4
Gulf Shellfish Institute, 13230 Eastern Ave, Palmetto, FL 34221, USA
5
The Whitney Laboratory for Marine Bioscience, 9505 N Ocean Shore Blvd, St. Augustine, FL 32080, USA
6
Institute of Food and Agricultural Sciences, Tropical Research and Education Center, University of Florida, Homestead, FL 33031, USA
7
St. Johns River Water Management District, 4049 Reid Street, Palatka, FL 32177, USA
*
Author to whom correspondence should be addressed.
J. Mar. Sci. Eng. 2024, 12(10), 1847; https://doi.org/10.3390/jmse12101847
Submission received: 26 July 2024 / Revised: 23 September 2024 / Accepted: 11 October 2024 / Published: 16 October 2024
(This article belongs to the Special Issue Coastal Ecological Restoration: Techniques and Novel Approaches to Living Shorelines and Oyster Reef Construction)
Browse Figures
Versions Notes

Abstract

:
Seagrasses provide valuable ecosystem services, including improved water quality, carbon sequestration, and sediment stabilization. Unfortunately, these critical habitats are declining globally due to a range of anthropogenic impacts. Restoration practitioners have made efforts to mitigate loss through the introduction of seagrass transplants. However, seagrass restoration has low success rates and is plagued by water quality concerns and ecosystem degradation. Studies to improve seagrass restoration efforts have targeted different functional taxa to allow for greater stability and recovery of threatened ecosystems, and excluded macro-grazers to limit losses to newly established and restored seagrasses. We hypothesized greater seagrass (Halodule wrightii) health when restored in conjunction with hard clams (Mercenaria mercenaria) and protected from grazers using herbivore exclusion devices (HEDs) in the Indian River Lagoon (IRL), Florida. While our study experienced high rates of seagrass mortality, we were able to observe significant differences in seagrass health between IRL sub-lagoons and observed the positive effects of HEDs on seagrass health. The observed high species mortality highlights how restoration in the IRL is hindered by biotic and abiotic stressors, site selection, and a lack of clear restoration protocols. As we see ongoing declines in water quality and loss of foundational species, informed site selection, greater understanding of grazer presence, and multi-species restoration will provide an informed approach for future seagrass restoration projects globally.

1. Introduction

Seagrasses are keystone estuarine and marine organisms, yet 14% of all seagrass species face extinction [1,2]. Seagrasses are sensitive to environmental conditions and their decline is a sentinel of change, which signals the loss of ecosystem services and detrimental shifts in ecosystem structure [3]. The global trajectory of seagrass decline was estimated at 7% annual loss from 1990 to 2009; however, losses are non-linear and highly variable [4,5]. Restoration initiatives have been prompted by these losses, but they are often highly unsuccessful with low survival rates [6,7]. Seagrass restoration occurs in dynamic environments challenged by climate change, predator interactions, physical disturbances, and increased anthropogenic activities, all contributing to the high failure rate of transplant success [3,8,9]. Reported restoration success is also skewed by short-term monitoring (<1 year), which shows early success that may not be sustained past the initial growing season [6]. Projects are difficult to compare, due to both a lack of monitoring and a lack of standardization in the monitoring that is undertaken [10].
The health of seagrasses, and thereby the success of restoration, can be improved by utilizing interspecific mutualism through co-restoration [11]. Mechanisms of positive species interactions between suspension feeding bivalves and sub-tropical seagrasses are equivocal [12,13]. The inclusion of bivalves can improve seagrass survival and growth, especially in areas recovering from environmental disturbance [11]. Hard clams (Mercenaria mercenaria) increase seagrass growth in eutrophic systems by reducing chlorophyll a (Chl a) and epiphyte loading, while increasing light penetration [14,15,16]. Additionally, suspension feeding bivalves enrich sediments with nutrients through the production of biodeposits, providing an important nutrient source for seagrass growth [14,15]. Numerous biodiverse faunal communities, including valuable fisheries, are supported by seagrasses [17].
A long-term goal of seagrass restoration is to re-establish natural, persistent food chains capable of supporting macro-grazers. However, short-term goals must focus on initial seagrass survival until a critical threshold for success is reached [9]. Newly restored seagrasses are susceptible to transplant stress, including desiccation, and are more vulnerable than fully established seagrass meadows to natural disturbances such as wave exposure and grazing [18]. Moderate grazing is a common natural disturbance; however, “overgrazing” occurs when grazing rates exceed seagrass growth rates [19]. Grazers remove substantial seagrass biomass, influencing seagrass meadow structure and distribution, decreasing daily seagrass leaf productivity, and even removing belowground biomass [20,21,22]. Excluding herbivores limits this grazing stress and positively impacts seagrass biomass and meadow structure [23,24].
The IRL is a shallow, restricted lagoon system [25] spanning 256 km along Florida’s subtropical East coast. It is divided into five distinct sub-lagoons: North IRL (NIRL), Central IRL (CIRL), Southern IRL (SIRL), Banana River, and Mosquito Lagoon [26]. These IRL sub-lagoons differ in hydrological flow and water residence time, due to tidal influences, freshwater flows, ocean inlets, wind, and rainfall [27]. Five ocean inlets punctuate the IRL, with considerable distance between them, resulting in high water residence times (~230 days in the NIRL; [28]). This study focused on sites in the CIRL and the NIRL only (no sites were established in the SIRL, Banana River, or the Mosquito lagoon), due to resource limitations. The CIRL is located within proximity to four of these oceanic inlets, resulting in a lower water residence time and more hospitable conditions for seagrass growth. Spatial and temporal exchange of freshwater input from watersheds results in fluctuating salinity, temperature, and light availability, which drive seagrass distribution and cover [29]. Seagrass species richness increases near the southern inlets and decreases in areas with higher water residence times [30].
In the IRL, ~19,000 ha or ~58% of seagrasses were lost between 2011 and 2019 [29], primarily in response to low light availability from intense phytoplankton blooms and persistent brown tides in 2011 and 2016–2018 [31,32]. Halodule wrightii is a dominant pioneer species occurring throughout the Caribbean, Gulf of Mexico, and Southern Atlantic, and is prominent within the IRL. We chose this species for restoration, as it tolerates a wider range of conditions compared to other seagrasses [29,33,34]. Halodule wrightii is a known food source for the West Indian manatee (Trichechus manatus) within the Indian River Lagoon, making it one of the main herbivores we aimed to exclude in this study [21].
Within the Indian River Lagoon (IRL), the hard clam (Mercenaria mercenaria) held historical, commercial, and cultural value prior to their population collapse, causing hard clams to become nearly non-existent in the IRL in the late twentieth century [35]. In 2019, the University of Florida Whitney Laboratory for Marine Bioscience (WLMB, St. Augustine, FL) collected a broodstock of clams from the IRL. These “super clams” were aged and found to have survived the devastating brown tides and hypoxic events that occurred in 2012 [36]. Their offspring likely inherited favorable traits of survivability and resiliency, and the addition of these offspring to seagrass plots could lead to multi-species restoration success.
Motivated by dramatic losses of seagrass and challenges to restoration, we explored how clam presence (and their ecosystem services) and herbivore absence impacts seagrass restoration success in the IRL. The following hypotheses provide valuable insights for restoration practitioners:
(i)
H. wrightii restored in conjunction with M. mercenaria would show increased survival and growth compared to H. wrightii restored without M. mercenaria.
(ii)
H. wrightii contained within herbivore exclusion devices (HEDs) would show increased survival and growth compared to H. wrightii without herbivory exclusion.
(iii)
H. wrightii planted in the CIRL would show increased survival and growth compared to H. wrightii planted in the NIRL, due to spatial gradients in water quality.

2. Materials and Methods

2.1. Site Location

We investigated our hypotheses within the IRL (28.61 to 27.86, −80.54 to −80.81 decimal degrees, Table 1; Figure 1). Sites (NIRL: N1, N2, N3, N4 and CIRL: C1, C2, C3, C4, C5) were shallow (0.6–1.6 m max. depth), sand dominated, nearshore waters with relatively low wave action and were devoid of existing seagrass but known locations of historical seagrass meadows [37]. Halodule wrightii fragments were harvested from adult plants grown in the inland seagrass nursery at Florida Oceanographic Society. Mercenaria mercenaria were obtained from the Whitney Laboratory for Marine Bioscience and were the offspring of the aforementioned “super clams”.
Sites were established in June 2022, and subsequent monitoring was undertaken at months 1, 2, and 4 (July, August, and October 2022, respectively, Table 1).
Hurricane Ian made landfall on the Gulf Coast of Florida as a Category 4 storm on 8 September 2022 [38]. Within Brevard County, where all field sites were located, rainfall totals of 36.1 cm and wind gusts of 130.4 km/h were documented [39]. Additionally, Hurricane Nicole made landfall as a Category 1 storm on the Atlantic Coast of Florida 63 days later (10 November 2022), depositing 15.3 cm of rain and with wind gusts of 115.9 km/h within Brevard County [39]. As a result of these events (and study species mortality), the sampling frequency varied among sites (N1, N2, N3, N4: July only; C2, C3, C4: July, August; C1, C5: July, August, October).

2.2. Experimental Design

Within each site, one (1) m2 plots were established in a factorial design (n = 3). Plots contained either
(a)
seagrass only (SG, restoration),
(b)
seagrass and clams (SG + Clam, co-restoration),
(c)
sand only (C, control to monitor sediment characteristics).
Within sites, treatments were replicated in caged or open plots to assess macrograzer exclusion (n = 9). Treatments were randomly distributed across 18 plots and uniformly spaced at least 0.76 m apart (Figure 2). The spacing of plots was dictated by the amount of usable seabed within the boundaries of private properties and avoiding manmade structures.
Clams were seeded at a consistent density (111 individuals m2) two hours prior to seagrass deployment. Seagrass restoration units, containing 16 shoots/mat (256 shoots m2) were created using 20 × 20 cm squares of agriculture grade burlap (340 gsm, untreated, 100% virgin natural jute fiber) and 0.83 mm floral wire [40]. Units were deployed in a grid (4 × 4 mat array), creating a 0.64 m2 seagrass plot. To facilitate root establishment and reduce wave action, units were secured to the sediment using 15 cm bamboo picks and a light layer of sediment. Experiments were terminated if sites reached 100% seagrass loss (NIRL: August 2022, CIRL: October 2022).
Macrograzer exclusion was achieved through 0.81 m tall cylindrical HEDs that protected a 1 m2 area. HEDs were constructed of 6.45 cm2 grid, 1.15 mm diameter PVC coated galvanized steel wire and 1.75 cm stainless steel fasteners. They were anchored to the sediment using three 2 m segments of rebar, oriented vertically and hammered 0.5 m into the sediment to create a supportive frame to apply the steel wire. Wire was wrapped around the rebar frame and a top was secured over the upright ends of the rebar. HED’s were cleaned weekly to remove epiphytic growth and biofouling.
Clams used in this study originated from the same size category at the aquaculture facility. However, to further assess baseline clam size, non-lethal measurements of body length, height, and width (mm) were taken for 450 clams from the same offspring cohort collected at clam deployment.
An additional 90 clams were haphazardly selected and processed to assess condition index (CI) using dry tissue weight: dry shell weight [41,42].

2.3. Seagrass and Hard Clams

In August and October, clam growth and survivorship were assessed by taking body measurements of live and dead clams in a 0.25 m2 quadrat within all plots. At the termination of the last two sites, a subsample of live clams (n = 16) was removed for CI analysis.
To test if the presence of clams or absence of herbivores increased seagrass health, seagrass percent cover (%) and canopy height (mm) were recorded during each sampling trip. A modified version of the protocols established in St. Johns River Water Management District (SJRWMD) Indian River Lagoon Seagrass Monitoring Standard Operating Procedures was used [43]. Seagrass coverage was evaluated using visual estimates of the percent cover within a quadrat divided by strings into cells, and height measurements were taken from sediment to leaf tip. Seagrass sampling methods were selected based on existing sampling methods used in the IRL and to best suit the environmental conditions at the sites.

2.4. Sediment

To determine potential site differences between the CIRL and NIRL, sediment characteristics were taken using 10 cm deep cores, which were collected using a 3 cm diameter syringe hand corer. Cores were stored at 4 °C, then dried at 60 °C until they reached a constant weight. Bulk density (g cm−3) was calculated as oven-dried mass divided by sample volume. Samples were ground for homogeneity and subsamples were combusted at 500 °C for 4 h in an Isotemp muffle furnace (Fisher Scientific) for soil organic matter (SOM) by loss-on-ignition. Remaining subsamples were combusted for total carbon (C) and total nitrogen (N) with a carbonate correction using an EA 1108 CHNS—O (Carlo Erba Instruments, Milan, Italy) at the University of Florida, Gainesville, USA. Grain size fractions (%) were calculated using the method outlined in Folk [44]. Baseline hydro-edaphic characteristics were measured in June (n = 9), and treatment measurements were taken in October (n = 2). Porewater for salinity and nutrient (NO3, NH4, TP) analyses was collected at 15 cm using a sipper. Porewater NH4 and NO3− nutrient analyses were filtered with a 0.45µ PTFE hydrophilic filter and preserved with 12N H2SO4. Samples for TP were not filtered or preserved prior to storage. Nutrient analyses were performed with a AQ400 Discrete Analyzer (Seal Analytical), following chemistries adapted from standard EPA methods for nutrient analyses, at the WLMB (St. Augustine, FL, USA).

2.5. Water Quality

Water quality data were obtained for 1 June–1 October 2022 from three metered stations in the Indian River Lagoon Observatory Network (IRLON) of Environmental Sensors [45] (Figure 1). Stations were selected based on proximity to field sites (all <18 km); Titusville (TS, 28.598500, −80.791300), Malabar (MB, 27.975100, −80.536800), and Sebastian (SB, 27.839100, −80.470800), MB and SB were pooled for CIRL. Photosynthetically active radiation (PAR; μmol m−2 s−1), chromophoric dissolved organic matter (CDOM; QSDE), chlorophyll (μg L−1), dissolved oxygen (mg L−1), salinity (PPT), turbidity (NTU), water temperature (°C), and pH were available at all stations and recorded by in situ IRLON instruments. Hourly data were downloaded with the IRLON quality filter turned on, to include only data that passed critical real-time quality control tests and deemed adequate for use as preliminary data. Daily means were calculated for all parameters excluding PAR, for which daylight averages were calculated using values between 08:00 and 16:00.
Additionally discrete water quality measurements were taken at every site during each sampling campaign, location of measurements was haphazardly chosen. pH, dissolved oxygen (%; mg L−1), salinity (ppt), and water temperature (°C) were measured with a calibrated multiparameter data sonde (YSI ProDSS 2020, Yellow Spring, OH, USA), and turbidity was measured using a LaMotte Turbidimeter 2020 (Chestertown, MD, USA).
To determine the effect of HEDs, water samples were collected from a subset of caged and uncaged plots (n = 3) to determine total suspended solids (TSS), particulate organic matter (POM), particulate inorganic matter (PIM), and chlorophyll a (Chl a) across treatments. Samples were kept cool and in dark bottles prior to lab processing. For TSSs, water was filtered in a pre-weighed 0.7 µm glass fiber filter (GF/F) and dried at 60 °C until a constant dry weight was achieved (12–48 h). TSSs were then obtained by re-weighing the filter and subtracting the initial filter weight. Dried filters were then combusted at 500 °C for one hour to burn off organic matter and re-weighed, POM was calculated using the weight lost after combustion, and was calculated PIM using the weight remaining on the filter. All methods were adapted from the Standard Methods for the Examination of Water and Wastewater [46].
To determine clam food availability, samples were analyzed for Chl a. Water column samples were filtered using 0.7 µm GF/F filters and the filter frozen until extraction. Filters were then ground with 10 mL of 90% acetone and allowed to extract for 2–24 h in the dark under refrigeration. After extraction, room temperature samples were centrifuged at ~2500 rpm for 20 min then measured fluorometrically using a Turner Designs Trilogy fluorometer (San Jose, CA, USA) before and after acidification using 10% HCl for Chl a concentration versus phaeophytin concentrations.
To determine the light availability for seagrass, PAR (μmol m−2 s−1) was measured using a Li-COR 1500 Light Sensor Unit with two calibrated LI-193 Spherical Quantum Sensors (Lincoln, NE, USA). PAR was recorded at the water surface and sediment surface, and calculated as the percentage of PAR that reached the sediment. These measurements were taken in all plots to test for differences between caged/uncaged treatments (n = 9). Discrete PAR measurements were taken during site sampling events, all within daylight hours of 09:00–14:00.

2.6. Statistical Analyses

All data were tested for normality using Shapiro–Wilks tests and distribution histograms.
Differences between baseline data discrete water quality parameters: turbidity (NTU), salinity (ppt), dissolved oxygen (DO, mg/L), pH, water temperature (°C), air temperature (°C), sediment bulk density, carbon or nitrogen, porewater NO3, chlorophyll a (µg/L), TP, and grain size were tested using Welch t-sample tests when normal distributions were present or a Wilcox rank test when there was not a normal distribution (Table S1).
Clam size measurement correlation was assessed using Kendall rank correlation tests. Differences between baseline clam lengths (mm) were tested using the nonparametric Kruskal–Wallis test, as there was not a normal distribution (Table S4).
For data taking during monitoring events, when data were not normal, generalized linear mixed-effects models (GLMs) using the ‘lme’ R package with three effects (sub-lagoon, seagrass treatment, and HED) and one random effect (plot) were used to determine significant differences for seagrass metrics (H. wrightii cover (%) and H. wrightii height) and PAR (% reached sediment) for the July sampling. GLMs were used for the August and October sampling, sub-lagoon was removed as an effect. Type II Wald chi-square tests using GLMs were performed to construct analysis of deviance tables, when significant factors or interactions were found, post hoc estimated marginal means were performed using the R package emmeans [estimated marginal means, also known as Least-Squares Means_. R package version 1.10.3] for pairwise comparisons. These tests were performed for August samplings of seagrass metrics, PAR, TSS, PIM, clam mortality, and Chl a; and October samplings of seagrass metrics, PIM, Chl a, total nitrogen, NH4, TP, porewater salinity, C:N ratio, and NO3.
When data were normal, analyses of variance (ANOVA) with two factors (seagrass treatment and HED) were run, correcting errors using plots. These tests were run for August sampling of POM (log transformed) and clam length; and October samplings of PAR, LOI, bulk density, TSS (log transformed), total carbon (square-root transformed), and POM (square-root transformed); Tables S1–S5.
All analyses were conducted with R (version 4.1.0 (18 May 2021)) and RStudio (2022.07.1 + 554, 2022.07.2 + 576, or 2022.12.0 + 353) with a significance level of alpha = 0.05.

3. Results

3.1. Seagrass Restoration

Restored seagrasses exhibited 100% mortality two-months post-deployment in the NIRL and 4-months post deployment in the CIRL.
An initial sampling event 1-month post deployment found seagrass cover was 96.75% higher in the CIRL than in the NIRL (X21,106 = 23.2036, p < 0.001, Figure 3).
Canopy height was 69.18% higher in the CIRL than in the NIRL and greater in caged than uncaged plots (X21,106 = 39.101, X21,106 = 10.544, both p < 0.001, Figure 3 and Figure 4, respectively).
A secondary sampling event 2 months post deployment was undertaken at CIRL sites with surviving seagrass populations (n = 5). Caged plots had 77.62% greater seagrass cover and 44.33% taller canopy height than uncaged plots in the CIRL (X21,59 = 7.5833, p < 0.05; X21,27 = 8.9182, p < 0.05, respectively, Figure 5).

3.2. Clam Introduction

Clam height was used as a proxy for clam size as it positively correlated with width and length at all sampling events (Kendall rank correlation: τb = 0.658, 0.807; τb = 0.899, 0.912; τb = 0.856, 0.906, respectively, all p < 0.001). There was no significant variation in clam length for each plot from baseline measurements (W1,49 = 49.76, p = 0.454). Clams showed very high rates of mortality. Clams in the NIRL reached 100% mortality 1 month post-deployment. In the CIRL, clams in uncaged plots reached 100% mortality 4 months post deployment, no differences were found in clam length or mortality between caged or uncaged plots (Tables S4 and S5).

3.3. Sediment and Water Quality

For baseline discrete water quality parameters, no significant differences were found between sub-lagoons for turbidity (NTU), salinity (ppt), dissolved oxygen (DO, mg L−1), pH, water temperature (°C), Chl a, air temperature (°C), sediment bulk density, C or N, or in porewater NO3, TP, or grain size (all p > 0.05, Table S1).
In June, the CIRL had significantly higher TSS, POM, PIM, sediment organic matter (LOI), CN, porewater salinity, and NH4 than the NIRL during baseline measurements (all p ≤ 0.05, Table S1).
In August, no significant differences were found in TSS, POM, PIM between caged and uncaged plots, (all p > 0.05), but caged plots had greater Chl than uncaged plots (X21,27 = 6.1549, p = 0.0131, Figure 5, Table S4).
In October, no significant differences were found in Chla, PIM, POM, CN, total sediment carbon, porewater salinity, NH4, NO3, or TP between caged and uncaged plots or seagrass treatments, all p > 0.05, Table S5).
Caged plots had greater total sediment nitrogen (%), lower bulk density and greater organic matter (LOI) than uncaged plots (X21,27 = 6.323, p = 0.012, F1,34 = 6.937, p = 0.022; F1,34 = 7.416, p = 0.019, respectively, Figure 5, Table S5). For bulk density, ANOVA found HED x treatment interactions, for results of post hoc estimated marginal means see Figure S1.
Caged plots had less PAR reaching sediment (%) than uncaged plots 1 month post deployment (X21,27 = 7.2683, p = 0.007, Figure 6a). Plots in the NIRL had less PAR reaching sediment (%) than CIRL plots 1 month post deployment (X21,27 = 4.3135, p = 0.038, Figure 6b, Table S3).
From IRLON water quality measurements, there were significant differences between sub-lagoons for CDOM (QSDE), chlorophyll (Chl; μg L−1), PAR (μmol m−2 s−1), salinity (ppt), turbidity (NTU), and water temperature (°C, Figure 7; Table S2).

4. Discussion

Our study highlights the fragility and variability in seagrass restoration in the IRL and shows success is driven by both wide-scale latitudinal environmental gradients and local site-specific parameters. Our preliminary results found seagrass growth and survival was higher in the CIRL and in plots with HEDs. However, two months post-planting seagrass mortality was 100% in the NIRL, and 50% in the CIRL. These drastic differences clearly show that seagrass restoration success varies based on location within the IRL and the exclusion of herbivores, highlighting the need for standardized site selection protocols and further studies documenting long-term outcomes of seagrass restoration in this region.

4.1. Water Quality

Significant differences were observed between sub-lagoon salinity and water clarity, which are crucial to seagrass growth and survival [47]. Salinity was higher, and turbidity and CDOM were lower, in the CIRL than in the NIRL (48%, 17%, and 41%, respectively). Halodule wrightii percent cover and canopy height were higher in the CIRL than the NIRL 1 month post-planting (69.18%, 96.75%; respectively).
Hydrology, freshwater inflow, and watershed land use all have a significant impact on water quality and therefore seagrass health in the IRL. Longer water residence times and poor flushing in the NIRL exposes seagrass populations to pollutants and terrestrial runoff, leading to poor water quality and substantial seagrass loss [29,48]. The impact of hurricane weather conditions, as experienced in this study, may have resulted in significant water quality changes due to increased runoff from rainfall.
Low salinity in the NIRL is likely a result of high rainfall, poor stormwater infrastructure, and limited tidal flushing [49]. Compared to other seagrasses, H. wrightii has a higher tolerance for low salinity; however, during our study, salinity in the NIRL reached levels considered detrimental to H. wrightii productivity for 44 days <20 ppt, [29,50,51]. Periods of prolonged low salinity can be particularly harmful to newly restored seagrasses already suffering from physical damage, desiccation, or impaired function due to transplantation stress [52].
Globally, turbidity, watercolor (CDOM), and Chl concentrations are responsible for attenuating downwelling light [53,54]. Within the IRL, poorly flushed and highly urbanized regions (e.g., the NIRL) [31] have low light availability, which is a major factor influencing seagrass distribution [55]. CDOM quality, composition, and concentration is heavily influenced by watershed land use [56] and differences in sub-lagoon CDOM are likely tied to urban spread and coastal development (CDOM was 41% lower in the CIRL than the NIRL). High Chl is associated with algae blooms from escalating eutrophication caused by coastal development, urbanization of watersheds, and septic system effluent in the IRL [31,57,58].
In the NIRL, shallow water, reduced ocean flushing, and limited flow result in increased temperatures [48]. During our study, mean sub-lagoon water temperatures fell within the established limits for healthy H. wrightii populations (~20–33 °C) [29,59]. However, increased water temperatures have also been associated with the prevalence of dinoflagellate phytoplankton biomass [60], which can block light and increase turbidity [61], proving detrimental to seagrass growth. Environmental conditions impair seagrass growth and survival, and improper site selection is a major cause of seagrass restoration failure [7,62]. Initial site assessments should consider wave action, natural disturbances, boating activity, light attenuation, sediment composition, and water depth [63,64,65]. We suggest that the water quality within the NIRL is not currently advantageous to the successful long-term growth and survival of H. wrightii, and seagrass restoration in this region is inadvisable until conditions improve [31].
Water quality is also crucial for clam growth and survival and M. mercenaria cannot tolerate low salinity environments for prolonged periods (<25 ppt) [66]. Low salinities impact clam feeding behavior and performance in the IRL [67] and may contribute to high clam mortality rates. Without considerable improvement in water quality, these challenges will remain, particularly in the NIRL.

4.2. Hard Clam Inclusion

A main objective was to observe co-restoration between H. wrightii and M. mercenaria; however, the high level of species mortality limited the power of our analyses. We observed extensive seagrass loss at 1-month sampling in the NIRL, which became 100% loss by 2 months. Inclusion of hard clams did not show an increase in seagrass health or survival in this study. As active benthic suspension feeders, clams push their siphon above the sediment to filter particles from the water column [68]. We postulate that this feeding behavior disrupted seagrass restoration units and root systems, highlighting the need for new planting methods of clam and seagrass species in co-restoration projects. Temporally spacing the restoration efforts to allow one population to become established prior to introducing another species may resolve this disruption; we recommend further studies to decide the order in which taxa are deployed. Clams can reduce H. wrightii productivity due to space competition and bioturbation, which can result in areas of localized disturbance [11,13,69].
The inadvertent recruitment of hard clam predators including crabs and whelks [70,71] and large mobile predators such as durophagous rays [72] may have contributed to negative co-restoration outcomes. Predators, including Florida crown conch (Melongena corona), mud crab (Scylla serrata), blue crab (Callinectes sapidus), and knobbed whelk (Busycon carica) were abundant in all treatment plots. The HEDs did not provide complete exclusion of these predators due to burrowing and the mesh size, which permitted some predators to enter. Many predators are offered the same protection and benefits from seagrass as hard clams and are more abundant in vegetated habitats [73]. Additionally, the presence of these predators can cause a reduction in clam growth rate, due to lower feeding rates [74,75], and the clams used in this study had not reached full adult size, which may have limited their capacity to reduce chlorophyll and epiphytes and enrich sediment.
Our results did not show a positive relationship between hard clams and seagrass health; however, we believe other factors influenced this (taxa choice, restoration methodology, environmental conditions). Therefore, we highly encourage continued evaluation of plant–bivalve interactions and the implementation of positive co-restoration in future restoration efforts [11,76]. Future restoration may explore alternative planting methods or utilization of alternative clam species including native Mercenaria campechiensis (Southern hard clams), which utilize seagrass as their habitat of choice [77].

4.3. Herbivore Exclusion

Herbivore exclusion devices aim to block herbivore access to seagrass, but their use and success is dependent on water flow, tidal impacts, and targeted exclusion of site-specific species. Globally, protection of seagrasses with HEDs increases seagrass leaf length, biomass, and structural complexity [78,79,80]. In restoration, excluding herbivores increases transplant survival by preventing direct consumption of seagrass [81]. Our results reflect this, with HEDs leading to 77.62% greater seagrass coverage and 44.33% greater canopy height. The successful exclusion of larger herbivores (e.g., green turtles and manatees) effectively reduced grazing on restored seagrass units; however, other organisms still gained access to plots. Within the HEDs, predatory gastropod species (including, M. corona) were abundant, and on occasion larger fish species and otters (Lontra canadensis) were present, leading to plot disturbance.
Our results confirm that the use of HEDs can be beneficial to seagrass growth, whether through the exclusion of herbivores or other cage effects is undetermined. The use of HEDs had many other side effects aside from the intended herbivore exclusion. We found significantly more organic matter and total nitrogen (%) within sediments of caged plots than in uncaged plots. This accumulation of high-N sediment could be attributed to the grain size of the sediment that was accumulated, or the entrapment and decay of drift algae biomass within the HEDs [82].
Within the HEDs, there was a visible buildup of floccular sediment and biomass, which may have contributed to the higher quantities of organic matter and explain the increased growth and productivity of H. wrightii in these plots. Even with weekly intensive cleaning of HEDs, large quantities of macroalgae built up and PAR was lower in caged plots compared to uncaged. Chl a (µg L−1) levels were 3.03 times higher in caged than uncaged plots, indicating higher levels of photosynthetic material within HEDs. Fouling on HEDs can decrease nutrient advection, leading to higher water column nutrient levels [83], which explains the increased Chl a. Decomposition of the macroalgae which accumulated on the HED’s may have also been a contributing factor to the increased Chl a.
The HEDs had a high initial cost (~$60/HED), and construction and deployment took significant manpower and time. HEDs may not always be appropriate, given the costs, maintenance requirements, and impacts of fouling [78,84]. While others did not observe caging side effects [78,85], the inclusion of partial cage treatments would allow definitive observations of caging artefacts beyond the intended herbivore exclusion [86,87].
Our HEDs experienced disturbance from wave action and storm activity, leading to structural damage and herbivore access to plots. Improved HED designs could utilize a thicker mesh gauge for stability or a sediment skirt to exclude burrowing predators. Higher perimeter fencing without a top may reduce light limitation, be more flexible when exposed to high wave action, and provide a larger exclusion area using fewer materials. This would allow for large scale seagrass planting, which has higher success rates than smaller restoration projects [9]. Prior to establishing restoration sites and designing HEDs, grazing pressure should be evaluated, local ordinances should be reviewed, and herbivore communities should be cataloged to customize HEDs to exclude specific local herbivore communities [24,79,88].

5. Conclusions

Despite high rates of seagrass mortality, our results provided valuable and novel insights into seagrass restoration efforts and our data can provide a useful tool for others. Our study highlighted the importance of informed site selection and the value of HEDs. However, we recommend further study into HED side effects, specifically the impact they may have on sediment composition through accumulation of organic matter. Our results highlighted potential issues with co-restoration of seagrasses and bivalves and indicated a need for future evaluations of alternative restoration practices. The inclusion of bivalves and the exclusion of predators requires further investigation, and the challenges of water quality and disturbance within the IRL must be overcome to ensure future seagrass restoration success.

Supplementary Materials

The following supporting information can be downloaded at https://www.mdpi.com/article/10.3390/jmse12101847/s1, Table S1: Mean ± SD surface water, sediment and porewater in June 2022/baseline data, Table S2: Mean ± SD of surface water data from IRLON stations, Table S3: July 2022 Sampling: Analysis of Deviance Table (Type II Wald chisquare tests), Table S4: August 2022 Sampling: GLM/ANOVA Analysis of Deviance Table (Type II Wald chisquare tests), Table S5: October 2022 Sampling: GLM + ANOVA Analysis of Deviance Table (Type II Wald chisquare tests), Table S6: Clam statistical analysis, plot = caged/uncaged, and Figure S1: Post-hoc estimated marginal means for HED:Treatment interaction for sediment bulk density in October.

Author Contributions

Conceptualization, V.A.M., M.K.G. and L.T.S.; methodology, V.A.M., M.K.G., S.M.C., T.Z.O., A.R.S. and L.T.S.; formal analysis, V.A.M., M.K.G., S.M.C. and L.T.S.; investigation, V.A.M., M.K.G., S.M.C. and L.T.S.; resources T.Z.O., A.R.S. and L.T.S.; data curation, V.A.M., M.K.G., S.M.C. and L.T.S.; writing—original draft preparation, V.A.M.; writing—review and editing V.A.M., M.K.G., S.M.C., T.Z.O., A.R.S. and L.T.S.; project administration, M.K.G. and L.T.S.; funding acquisition, L.T.S. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by St. Johns River Water Management District, grant number 36524.

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

The raw data supporting the conclusions of this article will be made available by the authors on request.

Acknowledgments

The authors would like to thank homeowners who allowed unabridged access to their property, members of the Restore of Shores Program (Brevard Zoo) for their collaboration, and M. Sullivan for clam aquaculture. We also thank C.A. Rooney, J. Brown, B. Furman, A. Roddenberry, T. Fridrich, A. Klingenberg, and T. Provoncha for field and lab assistance. Lastly, we thank members of the public who volunteered their time on this project, including those who attended seagrass restoration workshops and “clam buddies”.

Conflicts of Interest

The authors declare no conflicts of interest. The funders had no role in the design of the study; in the collection, analyses, or interpretation of data; in the writing of the manuscript; or in the decision to publish the results.

References

  1. Barbier, E.B.; Hacker, S.D.; Kennedy, C.; Koch, E.W.; Stier, A.C.; Silliman, B.R. The value of estuarine and coastal ecosystem services. Ecol. Monogr. 2011, 81, 169–193. [Google Scholar] [CrossRef]
  2. Short, F.T.; Polidoro, B.; Livingstone, S.R.; Carpenter, K.E.; Bandeira, S.; Bujang, J.S.; Calumpong, H.P.; Carruthers, T.J.B.; Coles, R.G.; Dennison, W.C.; et al. Extinction risk assessment of the world’s seagrass species. Biol. Conserv. 2011, 144, 1961–1971. [Google Scholar] [CrossRef]
  3. Orth, R.J.; Carruthers, T.J.B.; Dennison, W.C.; Duarte, C.M.; Fourqurean, J.W.; Heck Jr, K.L.; Hughes, A.R.; Kendrick, G.A.; Kenworthy, W.J.; Olyarnik, S.; et al. A global crisis for seagrass ecosystems. Bio. Sci. Mag. 2006, 56, 987–996. [Google Scholar] [CrossRef]
  4. Waycott, M.; Duarte, C.M.; Carruthers, T.J.B.; Orth, R.J.; Dennison, W.C.; Olyarnik, S.; Calladine, A.; Fourqurean, J.W.; Heck, K.L.; Hughes, A.R.; et al. Accelerating loss of seagrasses across the globe threatens coastal ecosystems. Proc. Natl. Acad. Sci. USA 2009, 106, 12377–12381. [Google Scholar] [CrossRef] [PubMed]
  5. Dunic, J.C.; Brown, C.J.; Connolly, R.M.; Turschwell, M.P.; Côté, I.M. Long-term declines and recovery of meadow area across the world’s seagrass bioregions. Glob. Change Biol. 2021, 27, 4096–4109. [Google Scholar] [CrossRef]
  6. Cunha, A.H.; Marbá, N.N.; van Katwijk, M.M.; Pickerell, C.; Henriques, M.; Bernard, G.; Ferreira, M.A.; Garcia, S.; Garmendia, J.M.; Manent, P. Changing paradigms in seagrass restoration. Restor. Ecol. 2012, 20, 427–430. [Google Scholar] [CrossRef]
  7. Bayraktarov, E.; Saunders, M.I.; Abdullah, S.; Mills, M.; Beher, J.; Possingham, H.P.; Mumby, P.J.; Lovelock, C.E. The cost and feasibility of marine coastal restoration. Ecol. Appl. 2016, 26, 1055–1074. [Google Scholar] [CrossRef]
  8. Duarte, C.M. The future of seagrass meadows. Environ. Conserv. 2002, 29, 192–206. [Google Scholar] [CrossRef]
  9. van Katwijk, M.M.; Thorhaug, A.; Marbà, N.; Orth, R.J.; Duarte, C.M.; Kendrick, G.A.; Althuizen, I.H.J.; Balestri, E.; Bernard, G.; Cambridge, M.L.; et al. Global analysis of seagrass restoration: The importance of large-scale planting. J. Appl. Ecol. 2016, 53, 567–578. [Google Scholar] [CrossRef]
  10. Rezek, R.J.; Furman, B.T.; Jung, R.P.; Hall, M.O.; Bell, S.S. Long-term performance of seagrass restoration projects in Florida, USA. Sci. Rep. 2019, 9, 15514. [Google Scholar] [CrossRef]
  11. Valdez, S.R.; Zhang, Y.S.; van der Heide, T.; Vanderklift, M.A.; Tarquinio, F.; Orth, R.J.; Silliman, B.R. Positive ecological interactions and the success of seagrass restoration. Front. Mar. Sci. 2020, 7, 91. [Google Scholar] [CrossRef]
  12. Zhang, Y.S.; Gittman, R.K.; Donaher, S.E.; Trackenberg, S.N.; van der Heide, T.; Silliman, B.R. Inclusion of intra- and interspecific facilitation expands the theoretical framework for seagrass restoration. Front. Mar. Sci. 2021, 8, 645673. [Google Scholar] [CrossRef]
  13. Donaher, S.E.; Baillie, C.J.; Smith, C.S.; Zhang, Y.S.; Albright, A.; Trackenberg, S.N.; Wellman, E.H.; Woodard, N.; Gittman, R.K. Bivalve facilitation mediates seagrass recovery from physical disturbance in a temperate estuary. Ecosphere 2021, 12, e03804. [Google Scholar] [CrossRef]
  14. Peterson, B.J.; Heck, K.L. Positive interactions between suspension-feeding bivalves and seagrass—A facultative mutualism. Mar. Ecol. Prog. Ser. 2001, 213, 143–155. [Google Scholar] [CrossRef]
  15. Wall, C.C.; Peterson, B.J.; Gobler, C.J. Facilitation of seagrass Zostera marina productivity by suspension-feeding bivalves. Mar Ecol. Prog. Ser. 2008, 357, 165–174. [Google Scholar] [CrossRef]
  16. Gobler, C.J.; Doall, M.H.; Peterson, B.J.; Young, C.S.; DeLaney, F.; Wallace, R.B.; Tomasetti, S.J.; Curtin, T.P.; Morrell, B.K.; Lamoureux, E.M.; et al. Rebuilding a collapsed bivalve population, restoring seagrass meadows, and eradicating harmful algal blooms in a temperate lagoon using spawner sanctuaries. Front. Mar. Sci. 2022, 9, 911731. [Google Scholar] [CrossRef]
  17. Unsworth, R.F.K.; Cullen-Unsworth, L.C. Biodiversity, ecosystem services, and the conservation of seagrass meadows. Coast. Conserv. 2014, 19, 95. [Google Scholar]
  18. Suykerbuyk, W.; Govers, L.L.; Bouma, T.J.; Giesen, W.B.J.T.; de Jong, D.J.; van de Voort, R.; Giesen, K.; Giesen, P.T.; van Katwijk, M.M. Unpredictability in seagrass restoration: Analysing the role of positive feedback and environmental stress on Zostera noltii transplants. J. Appl. Ecol. 2016, 53, 774–784. [Google Scholar] [CrossRef]
  19. Eklöf, J.S.; de la Torre-Castro, M.; Gullström, M.; Uku, J.; Muthiga, N.; Lyimo, T.; Bandeira, S.O. Sea urchin overgrazing of seagrasses: A review of current knowledge on causes, consequences, and management. Estuar. Coast. Shelf Sci. 2008, 79, 569–580. [Google Scholar] [CrossRef]
  20. Bourque, A.S.; Kenworthy, W.J.; Fourqurean, J.W. Impacts of physical disturbance on ecosystem structure in subtropical seagrass meadows. Mar. Ecol. Prog. Ser. 2015, 540, 27–41. [Google Scholar] [CrossRef]
  21. Lefebvre, L.W.; Reid, J.P.; Kenworthy, W.J.; Powell, J.A. Characterizing Manatee habitat use and seagrass grazing in Florida and Puerto Rico: Implications for conservation and management. Pac. Conserv. Biol. 1999, 5, 289–298. [Google Scholar] [CrossRef]
  22. Christianen, M.J.A.; Herman, P.M.J.; Bouma, T.J.; Lamers, L.P.M.; van Katwijk, M.M.; van der Heide, T.; Mumby, P.J.; Silliman, B.R.; Engelhard, S.L.; van de Kerk, M.; et al. Habitat collapse due to overgrazing threatens turtle conservation in marine protected areas. Proc. R. Soc. B Biol. Sci. 2014, 281, 20132890. [Google Scholar] [CrossRef] [PubMed]
  23. Fourqurean, J.W.; Manuel, S.; Coates, K.A.; Kenworthy, W.J.; Smith, S.R. Effects of excluding sea turtle herbivores from a seagrass bed: Overgrazing may have led to loss of seagrass meadows in Bermuda. Mar. Ecol. Prog. Ser. 2010, 419, 223–232. [Google Scholar] [CrossRef]
  24. Lefebvre, L.W.; Provancha, J.A.; Slone, D.H.; Kenworthy, W.J. Manatee grazing impacts on a mixed species seagrass bed. Mar. Ecol. Prog. Ser. 2017, 564, 29–45. [Google Scholar] [CrossRef]
  25. Kjerfve, B.; Magill, K.E. Geographic and hydrodynamic characteristics of shallow coastal lagoons. Mar. Geol. 1989, 88, 187–199. [Google Scholar] [CrossRef]
  26. Kamerosky, A.; Cho, H.J.; Morris, L. Monitoring of the 2011 Super Algal Bloom in Indian River Lagoon, FL, USA, Using MERIS. Remote Sens. 2015, 7, 1441–1460. [Google Scholar] [CrossRef]
  27. Phlips, E.J.; Badylak, S.; Youn, S.; Kelley, K. The occurrence of potentially toxic dinoflagellates and diatoms in a subtropical lagoon, the Indian River Lagoon, Florida, USA. Harmful Algae 2004, 3, 39–49. [Google Scholar] [CrossRef]
  28. Smith, N.P. Tidal and nontidal flushing of Florida’s Indian River Lagoon. Estuaries 1993, 16, 739–746. [Google Scholar] [CrossRef]
  29. Morris, L.; Hall, L.; Miller, J.; Lasi, M.; Chamberlain, R.; Virnstein, R.; Jacoby, C. Diversity and distribution of seagrasses as related to salinity, temperature, and availability of light in the Indian River Lagoon, Florida. Seagrass Ecosyst. Anal. 2021, 84, 119–137. [Google Scholar]
  30. Morris, L.J.; Hall, L.M.; Jacoby, C.A.; Chamberlain, R.H.; Hanisak, M.D.; Miller, J.D.; Virnstein, R.W. Seagrass in a Changing Estuary, the Indian River Lagoon, Florida, United States. Front. Mar. Sci. 2022, 8, 789818. [Google Scholar] [CrossRef]
  31. Lapointe, B.E.; Herren, L.W.; Brewton, R.A.; Alderman, P.K. Nutrient over-enrichment and light limitation of seagrass communities in the Indian River Lagoon, an urbanized subtropical estuary. Sci. Total Environ. 2020, 699, 134068. [Google Scholar] [CrossRef] [PubMed]
  32. Phlips, E.J.; Badylak, S.; Nelson, N.G.; Hall, L.M.; Jacoby, C.A.; Lasi, M.A.; Lockwood, J.C.; Miller, J.D. Cyclical Patterns and a Regime Shift in the Character of Phytoplankton Blooms in a Restricted Sub-Tropical Lagoon, Indian River Lagoon, Florida, United States. Front. Mar. Sci. 2021, 8, 730934. [Google Scholar] [CrossRef]
  33. Bertelli, C.M.; Creed, J.C.; Nuuttila, H.K.; Unsworth, R.K.F. The response of the seagrass Halodule wrightii Ascherson to environmental stressors. Estuar. Coast. Shelf Sci. 2020, 238, 106693. [Google Scholar] [CrossRef]
  34. Biber, P. Prolonged low salinity tolerance in Halodule wrightii Asch. Aquat. Bot. 2022, 178, 103498. [Google Scholar] [CrossRef]
  35. Arnold, W.; Marelli, D.; Parker, M.; Hoffman, P.; Frischer, M.; Scarpa, J. Enhancing hard clam (Mercenaria spp.) population density in the Indian River Lagoon, Florida: A comparison of strategies to maintain the commercial fishery. J. Shellfish Res. 2002, 21, 659–672. [Google Scholar]
  36. Osborne, T.Z.; Martindale, M.Q.; Nunez, J.M. Restoration of Clam Populations in the Indian River Lagoon for Water Quality Improvement 2019–2020; Final Report; Indian River Lagoon National Estuary Program; 2020; 22p, Available online: https://irlcouncil.sharepoint.com/:b:/s/IRLNEPPublic/Eanr94TqEMxIrA7xMt-AXR0BNKPzYaoiSonDOt8aRbxXwQ?e=DkQRzA (accessed on 10 October 2024).
  37. St Johns River Water Management District (SJRWMD) OpenData_Wetlands_Seagrass. Available online: https://data-floridaswater.opendata.arcgis.com/ (accessed on 11 August 2023).
  38. NOAA Hurricane Ian’s Path of Destruction. National Environmental Satellite, Data, and Information Service. 2022. Available online: https://www.nesdis.noaa.gov/news/hurricane-ians-path-of-destruction (accessed on 11 August 2023).
  39. National Weather Service Rainfall Totals from Ian Across East Central Florida; NOAA: Washington, DC, USA, 2022.
  40. MacDonnell, C.; Tiling, K.; Encomio, V.; van der Heide, T.; Teunis, M.; Lengkeek, W.; Didderen, K.; Bouma, T.J.; Inglett, P.W. Evaluating a novel biodegradable lattice structure for subtropical seagrass restoration. Aquat. Bot. 2022, 176, 103463. [Google Scholar] [CrossRef]
  41. Lucas, A.; Beninger, P.G. The use of physiological condition indices in marine bivalve aquaculture. Aquaculture 1985, 44, 187–200. [Google Scholar] [CrossRef]
  42. Zeng, Y.; Yang, H. Review of molluscan bivalve condition index calculations and application in Northern Quahogs Mercenaria mercenaria. Aquac. Res. 2021, 52, 23–36. [Google Scholar] [CrossRef]
  43. Hall, L.M.; Morris, L. Indian River Lagoon Seagrass Monitoring Standard Operating Procedures Modified for Brevard County. Available online: https://www.brevardfl.gov/docs/default-source/soirl/indian-river-lagoon-seagrass-monitoring-standard-operating-procedures.pdf?sfvrsn=985f238_1 (accessed on 10 October 2024).
  44. Folk, R.L. Petrology of Sedimentary Rocks; Hemphill Pub. Co.: Austin, TX, USA, 1980; ISBN 0-914696-14-9. [Google Scholar]
  45. IRLON (No Date). Irlon.org. Available online: https://irlon.org/ (accessed on 11 August 2023).
  46. APHA 2540 SOLIDS; Standard Methods for the Examination of Water and Wastewater. American Public Health Association: Washington, DC, USA, 2017.
  47. Dennison, W.C. Effects of light on seagrass photosynthesis, growth and depth distribution. Aquat. Bot. 1987, 27, 15–26. [Google Scholar] [CrossRef]
  48. Sigua, G.C.; Steward, J.S.; Tweedale, W.A. Water-Quality Monitoring and Biological Integrity Assessment in the Indian River Lagoon, Florida: Status, Trends, and Loadings (1988–1994). Environ. Manag. 2000, 25, 199–209. [Google Scholar] [CrossRef]
  49. Phlips, E.J.; Badylak, S.; Grosskopf, T. Factors Affecting the Abundance of Phytoplankton in a Restricted Subtropical Lagoon, the Indian River Lagoon, Florida, USA. Estuar. Coast. Shelf Sci. 2002, 55, 385–402. [Google Scholar] [CrossRef]
  50. Lirman, D.; Cropper, W.P. The influence of salinity on seagrass growth, survivorship, and distribution within Biscayne Bay, Florida: Field, experimental, and modeling studies. Estuaries 2003, 26, 131–141. [Google Scholar] [CrossRef]
  51. Mazzotti, F.J.; Pearlstine, L.G.; Chamberlain, R.; Barnes, T.; Chartier, K.; DeAngelis, D. Stressor Response Models for Seagrasses, Halodule wrightii and Thalassia testudinum; Final Report to the South Florida Water Management District and the U.S. Geological Survey; University of Florida, Florida Lauderdale Research and Education Center: Fort Lauderdale, FL, USA, 2007; p. 19. [Google Scholar]
  52. Li, W.-T.; Kim, Y.K.; Park, J.-I.; Zhang, X.; Du, G.-Y.; Lee, K.-S. Comparison of seasonal growth responses of Zostera marina transplants to determine the optimal transplant season for habitat restoration. Ecol. Eng. 2014, 71, 56–65. [Google Scholar] [CrossRef]
  53. Kahn, A.E.; Durako, M.J. Photophysiological responses of Halophila johnsonii to experimental hyposaline and hyper-CDOM conditions. J. Exp. Mar. Biol. Ecol. 2008, 367, 230–235. [Google Scholar] [CrossRef]
  54. Adams, M.P.; Hovey, R.K.; Hipsey, M.R.; Bruce, L.C.; Ghisalberti, M.; Lowe, R.J.; Gruber, R.K.; Ruiz-Montoya, L.; Maxwell, P.S.; Callaghan, D.P.; et al. Feedback between sediment and light for seagrass: Where is it important? Limnol. Oceanogr. 2016, 61, 1937–1955. [Google Scholar] [CrossRef]
  55. Steward, J.S.; Virnstein, R.W.; Lasi, M.A.; Morris, L.J.; Miller, J.D.; Hall, L.M.; Tweedale, W.A. The impacts of the 2004 hurricanes on hydrology, water quality, and seagrass in the Central Indian River Lagoon, Florida. Estuaries Coasts 2006, 29, 954–965. [Google Scholar] [CrossRef]
  56. Shao, T.; Wang, T. Effects of land use on the characteristics and composition of fluvial chromophoric dissolved organic matter (CDOM) in the Yiluo River Watershed, China. Ecol. Indic. 2020, 114, 106332. [Google Scholar] [CrossRef]
  57. Lapointe, B.E.; Herren, L.W.; Debortoli, D.D.; Vogel, M.A. Evidence of sewage-driven eutrophication and harmful algal blooms in Florida’s Indian River Lagoon. Harmful Algae 2015, 43, 82–102. [Google Scholar] [CrossRef]
  58. Herren, L.W.; Brewton, R.A.; Wilking, L.E.; Tarnowski, M.E.; Vogel, M.A.; Lapointe, B.E. Septic systems drive nutrient enrichment of groundwaters and eutrophication in the urbanized Indian River Lagoon, Florida. Mar. Pollut. Bull. 2021, 172, 112928. [Google Scholar] [CrossRef]
  59. Koch, M.S.; Schopmeyer, S.; Kyhn-Hansen, C.; Madden, C.J. Synergistic effects of high temperature and sulfide on tropical seagrass. J. Exp. Mar. Biol. Ecol. 2007, 341, 91–101. [Google Scholar] [CrossRef]
  60. Phlips, E.; Badylak, S.; Quinlan, E.; Cichra, M. Factors affecting the distribution of Pyrodinium bahamense var. bahamense in coastal waters of Florida. Mar. Ecol. Prog. Ser. 2006, 322, 99–115. [Google Scholar] [CrossRef]
  61. Schmidt, A.L.; Wysmyk, J.K.C.; Craig, S.E.; Lotze, H.K. Regional-scale effects of eutrophication on ecosystem structure and services of seagrass beds. Limnol. Oceanogr. 2012, 57, 1389–1402. [Google Scholar] [CrossRef]
  62. Fonseca, M.S. Addy Revisited: What Has Changed with Seagrass Restoration in 64 Years? Ecol. Restor. 2011, 29, 73–81. [Google Scholar] [CrossRef]
  63. Campbell, M. Getting the foundation right: A scientifically based management framework to aid in the planning and implementation of seagrass transplant efforts. Bull. Mar. Sci. 2002, 71, 1405–1414. [Google Scholar]
  64. van Katwijk, M.M.; Bos, A.R.; de Jonge, V.N.; Hanssen, L.S.A.M.; Hermus, D.C.R.; de Jong, D.J. Guidelines for seagrass restoration: Importance of habitat selection and donor population, spreading of risks, and ecosystem engineering effects. Mar. Pollut. Bull. 2009, 58, 179–188. [Google Scholar] [CrossRef]
  65. Hotaling-Hagan, A.; Swett, R.; Ellis, L.R.; Frazer, T.K. A spatial model to improve site selection for seagrass restoration in shallow boating environments. J. Environ. Manag. 2017, 186, 42–54. [Google Scholar] [CrossRef]
  66. Whetstone, J.M.; Sturmer, L.M.; Oesterling, M.J. Biology and culture of the hard clam (Mercenaria mercenaria); USDA, Southern Regional Aquaculture Center (SRAC) Publication No. 433; SRAC: Stoneville, MS, USA, 2005; p. 6. [Google Scholar]
  67. Galimany, E.; Lunt, J.; Freeman, C.J.; Reed, S.; Segura-García, I.; Paul, V.J. Feeding behavior of eastern oysters Crassostrea virginica and hard clams Mercenaria mercenaria in shallow estuaries. Mar. Ecol. Prog. Ser. 2017, 567, 125–137. [Google Scholar] [CrossRef]
  68. Judge, M.; Coen, L.; Heck, K. Does Mercenaria mercenaria encounter elevated food levels in seagrass beds? Results from a novel technique to collect suspended food resources. Mar. Ecol. Prog. Ser. 1993, 92, 141–150. [Google Scholar] [CrossRef]
  69. Rielly-Carroll, E.; Freestone, A.L. Habitat fragmentation differentially affects trophic levels and alters behavior in a multi-trophic marine system. Oecologia 2017, 183, 899–908. [Google Scholar] [CrossRef]
  70. Irlandi, E.A.; Peterson, C.H. Modification of animal habitat by large plants: Mechanisms by which seagrasses influence clam growth. Oecologia 1991, 87, 307–318. [Google Scholar] [CrossRef]
  71. Polyakov, O.; Kraeuter, J.N.; Hofmann, E.E.; Buckner, S.C.; Bricelj, V.M.; Powell, E.N.; Klinck, J.M. Benthic predators and northern quahog (=hard clam) (Mercenaria mercenaria Linnaeus, 1758) populations. J. Shellfish Res. 2007, 26, 995–1010. [Google Scholar] [CrossRef]
  72. Cahill, B.V.; McCulloch, K.L.; DeGroot, B.C.; Bassos-Hull, K.; Ajemian, M.J. Breaking bags and crunching clams: Assessing whitespotted eagle ray interactions with hard clam aquaculture gear. Aquacult. Environ. Interact. 2023, 15, 59–71. [Google Scholar] [CrossRef]
  73. Arnold, W.S.; Marelli, D.C.; Bert, T.M.; Jones, D.S.; Quitmyer, I.R. Habitat-specific growth of hard clams Mercenaria mercenaria (L.) from the Indian River, Florida. J. Exp. Mar. Biol. Ecol. 1991, 147, 245–265. [Google Scholar] [CrossRef]
  74. Nakaoka, M. Nonlethal Effects of Predators on Prey Populations: Predator-Mediated Change in Bivalve Growth. Ecology 2000, 81, 1031–1045. [Google Scholar] [CrossRef]
  75. Tomiyama, T. Lethal and non-lethal effects of an invasive naticid gastropod on the production of a native clam. Biol. Invasions 2018, 20, 2005–2014. [Google Scholar] [CrossRef]
  76. Gagnon, K.; Rinde, E.; Bengil, E.G.T.; Carugati, L.; Christianen, M.J.A.; Danovaro, R.; Gambi, C.; Govers, L.L.; Kipson, S.; Meysick, L.; et al. Facilitating foundation species: The potential for plant–bivalve interactions to improve habitat restoration success. J. Appl. Ecol. 2020, 57, 1161–1179. [Google Scholar] [CrossRef]
  77. Arnold, W.S.; Geiger, S.P.; Peters Stephenson, S. Mercenaria mercenaria introductions into Florida, USA, waters: Duration, not size of introduction, influences genetic outcomes. Aquat. Biol. 2009, 5, 49–62. [Google Scholar] [CrossRef]
  78. Bourque, A.S.; Fourqurean, J.W. Variability in herbivory in subtropical seagrass ecosystems and implications for seagrass transplanting. J. Exp. Mar. Biol. Ecol. 2013, 445, 29–37. [Google Scholar] [CrossRef]
  79. Burkholder, D.A.; Heithaus, M.R.; Fourqurean, J.W.; Wirsing, A.; Dill, L.M. Patterns of top-down control in a seagrass ecosystem: Could a roving apex predator induce a behaviour-mediated trophic cascade? J. Anim. Ecol. 2013, 82, 1192–1202. [Google Scholar] [CrossRef]
  80. Samper-Villarreal, J.; Moya-Ramírez, J.; Cortés, J. Megaherbivore exclusion led to more complex seagrass canopies and increased biomass and sediment Corg pools in a tropical meadow. Front. Mar. Sci. 2022, 9, 945783. [Google Scholar] [CrossRef]
  81. Wendländer, N.S.; Lange, T.; Connolly, R.M.; Kristensen, E.; Pearson, R.M.; Valdemarsen, T.; Flindt, M.R. Assessing methods for restoring seagrass (Zostera muelleri) in Australia’s subtropical waters. Mar. Freshw. Res. 2020, 71, 996–1005. [Google Scholar] [CrossRef]
  82. Veenstra, J.; Southwell, M.; Dix, N.; Marcum, P.; Jackson, J.; Burns, C.; Herbert, C.; Kemper, A. High carbon accumulation rates in sediment adjacent to constructed oyster reefs, Northeast Florida, USA. J. Coast. Conserv. 2021, 25, 40. [Google Scholar] [CrossRef]
  83. Moksnes, P.-O.; Gullström, M.; Tryman, K.; Baden, S. Trophic cascades in a temperate seagrass community. Oikos 2008, 117, 763–777. [Google Scholar] [CrossRef]
  84. Hanisak, M.D.; Virnstein, R.W. Evaluating the Feasibility of Transplanting to Promote Seagrass Recovery in the Indian River Lagoon; Harbor Branch Oceanographic Institute at Florida Atlantic University: Fort Pierce, FL, USA, 2015. [Google Scholar]
  85. Cheng, B.S.; Ruiz, G.M.; Altieri, A.H.; Torchin, M.E. The biogeography of invasion in tropical and temperate seagrass beds: Testing interactive effects of predation and propagule pressure. Divers. Distrib. 2019, 25, 285–297. [Google Scholar] [CrossRef]
  86. Campbell, J.E.; Altieri, A.H.; Johnston, L.N.; Kuempel, C.D.; Paperno, R.; Paul, V.J.; Duffy, J.E. Herbivore community determines the magnitude and mechanism of nutrient effects on subtropical and tropical seagrasses. J. Ecol. 2018, 106, 401–412. [Google Scholar] [CrossRef]
  87. Rodriguez, A.R.; Marco-Méndez, C.; Campbell, J.; Heck, K.L. Effects of varying types and amounts of herbivory and nutrient enrichment on a tropicalizing seagrass meadow. Front. Mar. Sci. 2022, 9, 892219. [Google Scholar] [CrossRef]
  88. Bessey, C.; Heithaus, M.R.; Fourqurean, J.W.; Gastrich, K.R.; Burkholder, D.A. Importance of teleost macrograzers to seagrass composition in a subtropical ecosystem with abundant populations of megagrazers and predators. Mar. Ecol. Prog. Ser. 2016, 553, 81–92. [Google Scholar] [CrossRef]
Jmse 12 01847 g001
Figure 1. Map of the IRL showing North IRL (NIRL, yellow), Central IRL (CIRL, green), NIRL sites (black squares), CIRL (black circles), and Indian River Lagoon Observatory Network (IRLON) stations (pink triangles). Map created using ArcGIS Pro software by © Esri, 2023).
Figure 1. Map of the IRL showing North IRL (NIRL, yellow), Central IRL (CIRL, green), NIRL sites (black squares), CIRL (black circles), and Indian River Lagoon Observatory Network (IRLON) stations (pink triangles). Map created using ArcGIS Pro software by © Esri, 2023).
Jmse 12 01847 g001
Jmse 12 01847 g002
Figure 2. Example of a site plan includes 18 plots, plots were 1 m2 and spaced at least 0.76 m apart. The yellow square labeled “C” indicates sand only control, green square labeled “SG” indicates seagrass only restoration, and pink square labeled “SG + Clam” indicates seagrass and clam co-restoration (n = 3 for all plot types). Dashed black outline indicates plots protected from herbivory by HED (n = 9). Plot location was randomized for each site and orientation of the grid within the sites was adapted to the space available.
Figure 2. Example of a site plan includes 18 plots, plots were 1 m2 and spaced at least 0.76 m apart. The yellow square labeled “C” indicates sand only control, green square labeled “SG” indicates seagrass only restoration, and pink square labeled “SG + Clam” indicates seagrass and clam co-restoration (n = 3 for all plot types). Dashed black outline indicates plots protected from herbivory by HED (n = 9). Plot location was randomized for each site and orientation of the grid within the sites was adapted to the space available.
Jmse 12 01847 g002
Jmse 12 01847 g003
Figure 3. Across all sites and treatments in July 2022, (a) Halodule wrightii cover (%) was 69.18% higher in CIRL than NIRL and (b) H. wrightii canopy height (mm) was 96.75% higher in CIRL than the NIRL. North sub-lagoon (N, n = 4) and Central sub-lagoon (C, n = 5), all p < 0.001 (Table S3). Outliers denoted by open circles, significance shown as asterisk, and error bars by dashed vertical lines.
Figure 3. Across all sites and treatments in July 2022, (a) Halodule wrightii cover (%) was 69.18% higher in CIRL than NIRL and (b) H. wrightii canopy height (mm) was 96.75% higher in CIRL than the NIRL. North sub-lagoon (N, n = 4) and Central sub-lagoon (C, n = 5), all p < 0.001 (Table S3). Outliers denoted by open circles, significance shown as asterisk, and error bars by dashed vertical lines.
Jmse 12 01847 g003
Jmse 12 01847 g004
Figure 4. Halodule wrightii height (mm) for plots with (caged) and without (uncaged) HEDs in July 2022 (n = 5). Greater seagrass height (mm) was observed in caged plots in the central sub-lagoon (C). Outliers denoted by open circles, significance shown as asterisk, and error bars by dashed vertical lines.
Figure 4. Halodule wrightii height (mm) for plots with (caged) and without (uncaged) HEDs in July 2022 (n = 5). Greater seagrass height (mm) was observed in caged plots in the central sub-lagoon (C). Outliers denoted by open circles, significance shown as asterisk, and error bars by dashed vertical lines.
Jmse 12 01847 g004
Jmse 12 01847 g005
Figure 5. (a) Halodule wrightii coverage (%), (b) H. wrightii canopy height (mm) and (c) Chl a (chlorophyll a) (µg L−1) for caged and uncaged plots from CIRL sites (n = 5) in August 2022, (d) total nitrogen (%), (e) organic matter (LOI), and (f) bulk density (g cm−3) for caged and uncaged plots from sites (n = 4) in October 2022. Combined results of plots with (SG + Clam), and without clams (SG). Outliers denoted by open circles, significance shown as asterisk, and error bars by dashed vertical lines.
Figure 5. (a) Halodule wrightii coverage (%), (b) H. wrightii canopy height (mm) and (c) Chl a (chlorophyll a) (µg L−1) for caged and uncaged plots from CIRL sites (n = 5) in August 2022, (d) total nitrogen (%), (e) organic matter (LOI), and (f) bulk density (g cm−3) for caged and uncaged plots from sites (n = 4) in October 2022. Combined results of plots with (SG + Clam), and without clams (SG). Outliers denoted by open circles, significance shown as asterisk, and error bars by dashed vertical lines.
Jmse 12 01847 g005
Jmse 12 01847 g006
Figure 6. Across all sites and treatments in July 2022, (a) PAR reaching the sediment (%) was 29.7% higher in caged than uncaged plots across both sub-lagoons and (b) PAR reaching the sediment (%) was 26.3% higher in the central sub-lagoon (n = 5) than the north sub-lagoon (n = 4, all p < 0.05, Table S3). Outliers denoted by open circles, significance shown as asterisk, and error bars by dashed vertical lines.
Figure 6. Across all sites and treatments in July 2022, (a) PAR reaching the sediment (%) was 29.7% higher in caged than uncaged plots across both sub-lagoons and (b) PAR reaching the sediment (%) was 26.3% higher in the central sub-lagoon (n = 5) than the north sub-lagoon (n = 4, all p < 0.05, Table S3). Outliers denoted by open circles, significance shown as asterisk, and error bars by dashed vertical lines.
Jmse 12 01847 g006
Jmse 12 01847 g007
Figure 7. IRLON water quality parameters from June to October 2022, North IRL (NIRL, n = 1, yellow) and Central IRL (CIRL, n = 2, green). In the CIRL, (a) CDOM was 41% lower, (b) Chl was 45% higher, (c) PAR was 16% higher, (d) salinity was 48% higher, (e) turbidity was 17% lower, and (f) water temperature was 0.65% lower than the NIRL (all Wilcoxon texts, p < 0.001).
Figure 7. IRLON water quality parameters from June to October 2022, North IRL (NIRL, n = 1, yellow) and Central IRL (CIRL, n = 2, green). In the CIRL, (a) CDOM was 41% lower, (b) Chl was 45% higher, (c) PAR was 16% higher, (d) salinity was 48% higher, (e) turbidity was 17% lower, and (f) water temperature was 0.65% lower than the NIRL (all Wilcoxon texts, p < 0.001).
Jmse 12 01847 g007
Table 1. Site number, IRL sub-lagoon, coordinates, date of site establishment (beginning) and ending, and number of sampling events undertaken.
Table 1. Site number, IRL sub-lagoon, coordinates, date of site establishment (beginning) and ending, and number of sampling events undertaken.
IRL
Sub-Lagoon		SiteLocation
(Decimal Degrees)		BeginningEndingSampling Events Undertaken
North (NIRL)N128.605383, −80.8050836 June 202215 August 20221
N228.490233, −80.77320029 June 202215 August 20221
N328.463771, −80.76162130 June 202216 August 20221
N428.605383, −80.8050838 June 202215 August 20221
Central (CIRL)C128.064207, −80.56440528 June 202219 October 20223
C228.058471, −80.5650318 June 202217 October 20222
C328.024883, −80.5415009 June 202217 October 20222
C428.015683, −80.5665337 June 202217 October 20222
C527.864933, −80.49268327 June 202217 November 20223
Disclaimer/Publisher’s Note: The statements, opinions and data contained in all publications are solely those of the individual author(s) and contributor(s) and not of MDPI and/or the editor(s). MDPI and/or the editor(s) disclaim responsibility for any injury to people or property resulting from any ideas, methods, instructions or products referred to in the content.

Share and Cite

MDPI and ACS Style

Main, V.A.; Gilligan, M.K.; Cole, S.M.; Osborne, T.Z.; Smyth, A.R.; Simpson, L.T. Challenges to Seagrass Restoration in the Indian River Lagoon, Florida. J. Mar. Sci. Eng. 2024, 12, 1847. https://doi.org/10.3390/jmse12101847

AMA Style

Main VA, Gilligan MK, Cole SM, Osborne TZ, Smyth AR, Simpson LT. Challenges to Seagrass Restoration in the Indian River Lagoon, Florida. Journal of Marine Science and Engineering. 2024; 12(10):1847. https://doi.org/10.3390/jmse12101847

Chicago/Turabian Style

Main, Vivienne A., Morgan K. Gilligan, Sarah M. Cole, Todd Z. Osborne, Ashley R. Smyth, and Loraé T. Simpson. 2024. "Challenges to Seagrass Restoration in the Indian River Lagoon, Florida" Journal of Marine Science and Engineering 12, no. 10: 1847. https://doi.org/10.3390/jmse12101847

APA Style

Main, V. A., Gilligan, M. K., Cole, S. M., Osborne, T. Z., Smyth, A. R., & Simpson, L. T. (2024). Challenges to Seagrass Restoration in the Indian River Lagoon, Florida. Journal of Marine Science and Engineering, 12(10), 1847. https://doi.org/10.3390/jmse12101847

Note that from the first issue of 2016, this journal uses article numbers instead of page numbers. See further details here.

Article Metrics

Back to TopTop