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diversity Article Coral Restoration Eectiveness: Multiregional Snapshots of the Long-Term Responses of Coral Assemblages to Restoration 1 , 2 , 3 1 1 Margaux Y. Hein * , Roger Beeden , Alastair Birtles , Naomi M. Gardiner , 4 5 6 7 8 Thomas Le Berre , Jessica Levy , Nadine Marshall , Chad M. Scott , Lisa Terry and 1 , 2 Bette L. Willis College of Science and Engineering, James Cook University, Townsville, QLD 4811, Australia; firstname.lastname@example.org (A.B.); email@example.com (N.M.G.); firstname.lastname@example.org (B.L.W.) Australian Research Council (ARC) Centre of Excellence for Coral Reef Studies, Townsville, QLD 4811, Australia Great Barrier Reef Marine Park Authority, Townsville, QLD 4811, Australia; email@example.com Reefscapers Pvt Ltd., Seamarc, Marine Discovery Centre, Landaa Giraavaru 20215, Maldives; firstname.lastname@example.org Coral Restoration Foundation, Key Largo, FL 03037, USA; email@example.com CSIRO Land and Water, ATSIP Building#145, James Cook University, Townsville, QLD 4811, Australia; firstname.lastname@example.org Conservation Diver, Evergreen, CO 80439, USA; email@example.com The Nature Conservancy Caribbean Program, 3052 Estate Little Princess, Christiansted, VI 00820, USA; firstname.lastname@example.org * Correspondence: email@example.com Received: 9 March 2020; Accepted: 27 March 2020; Published: 17 April 2020 Abstract: Coral restoration is rapidly becoming a mainstream strategic reef management response to address dramatic declines in coral cover worldwide. Restoration success can be deﬁned as enhanced reef functions leading to improved ecosystem services, with multiple beneﬁts at socio-ecological scales. However, there is often a mismatch between the objectives of coral restoration programs and the metrics used to assess their eectiveness. In particular, the scales of ecological beneﬁts currently assessed are typically limited in both time and space, often being limited to short-term monitoring of the growth and survival of transplanted corals. In this paper, we explore reef-scale responses of coral assemblages to restoration practices applied in four well-established coral restoration programs. We found that hard coral cover and structural complexity were consistently greater at restored compared to unrestored (degraded) sites. However, patterns in coral diversity, coral recruitment, and coral health among restored, unrestored, and reference sites varied across locations, highlighting dierences in methodologies among restoration programs. Altogether, dierences in program objectives, methodologies, and the state of nearby coral communities were key drivers of variability in the responses of coral assemblages to restoration. The framework presented here provides guidance to improve qualitative and quantitative assessments of coral restoration eorts and can be applied to further understanding of the role of restoration within resilience-based reef management. Keywords: coral assemblages; coral restoration; eectiveness; monitoring 1. Introduction The number of coral restoration programs is burgeoning in most reef regions in response to worldwide declines in coral cover in recent years [1–3]. Common objectives of these programs are Diversity 2020, 12, 153; doi:10.3390/d12040153 www.mdpi.com/journal/diversity Diversity 2020, 12, 153 2 of 22 to assist the recovery of reefs, protect endangered coral species, promote sustainable alternative livelihoods, and enhance conservation stewardship , but there is a general mismatch between the stated objectives of these programs and indicators used to assess their eectiveness. In general, most assessments of coral restoration eectiveness are largely focused on the number of ramets created, growth, and survival post-transplantation . A recent review of coral restoration eorts globally revealed a lack of appropriate and standardized monitoring of outcomes, with too short timeframes (median monitoring time of 12 months) to assess the potential of using restoration as a tool for resilience-based management . This lack of long-term comprehensive assessments of coral restoration eectiveness is widely criticized [6,7] and hinders the uptake of coral restoration within multi-scale resilience-based management frameworks . In addition, many studies are focused on site- or region-speciﬁc restoration programs [9,10], which has made comparative studies dicult and limited the development of broad best-practice recommendations. Improved resilience of degraded reefs is the ultimate objective of many coral restoration programs. Not only has “managing for reef resilience” become a major focus of reef management [8,11], but “re-establishing a self-sustaining, functioning coral reef ecosystem after a disturbance” is also the most commonly stated objective for coral restoration . However, measuring the resilience of an ecosystem is a dicult exercise that requires a range of metrics accounting for aspects of both recovery and resistance over time . Reef attributes like hard coral cover, species diversity, and structural complexity are directly related to reef resilience [8,11,13] and may be enhanced by restoration programs. Percent hard coral cover is the most widely used metric to document reef recovery (e.g., ), although its use in isolation has limited value [11,15,16]. At restoration sites, increased hard coral cover may prevent phase shifts to algal-dominated systems , enhance the recruitment of juvenile corals to damaged areas , as well as regenerate the structural complexity of a degraded reef . Structural complexity may also be increased directly by artiﬁcial structures used as surfaces for coral transplants. High structural complexity of reef systems has been shown to decrease the sensitivity of local coral assemblages to extreme weather events , and also improve reef recovery post-bleaching . Increased coral diversity on restored reefs may lead to increased biodiversity of associated vertebrates and invertebrates, and hence increased functional diversity present within the reef community . Increased functional diversity may increase the resistance of the reef community by expanding the range of its potential responses to disturbances . Assessing the potential for reef restoration to improve reef resilience thus necessitates looking at processes occurring at the scale of the benthic community rather than solely at the scale of coral fragments transplanted to a degraded reef. A set of six ecological indicators that could be used to characterize the resilience of a reef community, based on an evaluation of indicators used in terrestrial restoration (e.g., ) and reef resilience studies (e.g., ), were developed by . These are: (1) Coral diversity, (2) herbivore biomass and diversity, (3) benthic cover, (4) recruitment, (5) coral health, and (6) structural complexity. Although subsets of these indicators have been used to characterize the resilience of reef communities [11,13], a collective set of these indicators has not yet been applied to assessing the outcomes of a coral restoration program. While the capacity of a coral restoration program to aect one or more of these indicators positively is likely to be constrained by factors, such as the degradation state of the reef area to be restored or the types of strategies used to restore the coral community, in combination, they provide a holistic assessment of restoration eectiveness. The objectives and methodologies of coral restoration programs typically dier among reef regions. In the Caribbean, most programs aim to restore two critically endangered species of Acropora (e.g., ), while programs in the Indo-Paciﬁc are more focused on restoring reef structure and resilience . Many programs depend on the capacity of corals to reproduce asexually and use either fragments from donor colonies or fragments of opportunity. While each methodology has its strengths and limitations in diering contexts, there is a critical need to further our understanding of how these dierent methodologies impact the resilience of restored reef areas in the long term to better inform reef managers. Diversity 2020, 12, 153 3 of 22 The aim of this paper was to capture a snapshot of the responses of coral assemblages to long-term restoration practices at four locations with well-established coral restoration programs that dier in objectives, methodologies, and socio-cultural settings. At each reef location, ﬁve coral-based indicators of reef resilience were characterized: Coral cover, structural complexity, coral diversity, coral juveniles, and coral health. These ﬁve indicators of restoration eectiveness were then qualitatively compared among the four restoration programs to gain insights into how dierent restoration designs inﬂuence the response of coral assemblages to coral restoration. The response of ﬁsh assemblages to these coral restoration eorts were examined independently. 2. Materials and Methods 2.1. Study Sites The four restoration programs selected had been in operation for 8 to 12 years, enabling assessments of the long-term eectiveness of diering restoration approaches. The programs selected represented four reef regions: (1) New Heaven Reef Conservation Program (NHRCP) on the island of Koh Tao, Thailand; (2) Reefscapers program on the island of Landaa Giraavaru, Maldives; (3) Coral Restoration Foundation in Key Largo, Florida Keys, USA; and (4) The Nature Conservancy on the island of St Croix, US Virgin Islands (Figure 1). Each location has a unique history of reef-associated disturbances; therefore, objectives for coral restoration varied from growing and restoring endangered species of corals (Florida Keys and St Croix), to restoring coral abundance and diversity at sites degraded by tourism pressures and bleaching events (Koh Tao and Landaa Giraavaru). Programs also diered in the set of coral restoration techniques used (see boxes 1–4; summarized in Figures 1 and 2), which provided an opportunity to qualitatively compare the relative eectiveness of dierent methodologies across the ﬁve indicators of reef resilience. Figure 1. Map showing the locations of the four coral restoration programs surveyed and an overview of the restoration strategies used in each program (see key at bottom of ﬁgure to interpret diagrams that represent techniques present at each site). Half green and half blue circles indicate adjacent restored and unrestored sites; red circles indicate control reference sites. Diversity 2020, 12, 153 4 of 22 Figure 2. Photo montage illustrating coral restoration strategies at the four coral restoration programs surveyed. Photo credits to Margaux Hein, New Heaven Reef Conservation Program, Reefscapers and Marine Savers, and The Coral Restoration Foundation. 2.1.1. Box 1. Coral Restoration in Koh Tao, Thailand Koh Tao is a moderately sized, high island (21 km in area) located in the Gulf of Thailand. The island has undergone rapid development in the past 30 years and is now considered a global hotspot for SCUBA diving, with over 500,000 visitors every year . This rapid development has been largely unregulated, and resorts, bars, and restaurants have replaced primary forests. What were once some of Thailand’s most biodiverse and pristine reefs are now under stress from terrestrial run-o and sedimentation [25,26], over-use by the local water-based tourism industry [27,28], and both land-based and marine pollution . Several studies have documented a high prevalence of coral disease and other indicators of compromised health [30–32]. Mass bleaching events recorded in 1998, 2010, and 2014 have also caused substantial coral mortality . A restoration program led by the New Heaven Reef Conservation Program (NHRCP) was initiated in 2007 to assist the recovery of locally degraded reefs by re-building the complexity of coral assemblages, increasing coral cover, and alleviating diving pressures through widespread education. NHRCP uses a wide range of coral restoration techniques, from direct transplantation of coral fragments into natural holes and crevices on the reef to the building of artiﬁcial reef structures. Artiﬁcial structures are used preferentially in areas where the reef structure has been compromised by boat groundings and anchors, or smothered by sediment run-o from adjacent land. Types of structures used include steel frames, electriﬁed artiﬁcial reefs, concrete reef balls, and glass bottles embedded in concrete (Figure 2A). Corals are collected as fragments of opportunity, attached to mid-water ropes and table nurseries (Figure 2A) for a few months, and then attached onto the reef or onto one of the artiﬁcial structures. Attachment methods vary from epoxy cement to nylon thread, cable ties, or ﬁne metal wire, depending on the type of structure. Restored areas are scattered around the island, and most include transplants attached to a variety of artiﬁcial reef structures, as well as directly onto the reef (Figure 1). Diversity 2020, 12, 153 5 of 22 2.1.2. Box 2. Coral Restoration in Landaa Giraavaru, Maldives Landaa Giraavaru is a small sand cay (0.18 km in area) situated in Baa Atoll, a UNESCO Biosphere Reserve since 2011, on the western front of the Maldivian atoll chain. One ﬁve-star luxury resort, comprised of 23 individual villas, was built in 2004 and occupies the whole cay. Construction of the resort caused substantial structural damage to local reefs, which also suered mass coral bleaching episodes and widespread coral mortality in 1998 and 2010 [34,35]. Coral restoration eorts led by the Reefscapers group (Reefscapers Pvt Ltd) primarily aim to increase biodiversity, reef complexity, and habitat diversity on the “house reef” surrounding the island. They use sand-coated steel structures, referred to as “coral frames”, as artiﬁcial substrata on which to attach coral fragments. Three sizes of frames are used (small, medium, and large), ranging from 110 40 cm to 200 110 cm (width height) (Figure 2B). Coral fragments are securely attached to frames with cable ties on land and the frames are then placed on the reef at depths ranging from 5 to 10 m around the island. As of March 2016, the reef around Landaa Giraavaru hosted 2800 frames, which covered an area of about 5500 m and harbored 40 dierent species of corals (Figure 1). The ﬁrst frames were populated with corals that were salvaged from the construction site when the resort was built in 2004. Nowadays, coral fragments are collected from colonies living on older frames, speciﬁcally targeting colonies that resisted earlier bleaching events. 2.1.3. Box 3. Coral Restoration in the Florida Keys, USA The Florida Keys in the United States of America have a long history of disturbances that have resulted in dramatic loss of coral cover and diversity, particularly in the past 20 years [1,36,37]. Disturbances have included tropical storms (2005, 2008, 2012), coral bleaching associated with both cold-water anomalies (2010) and warm water anomalies (1987, 1997, 1998, 2005, 2014), and severe outbreaks of coral disease and of corallivores [37–40]. Like Koh Tao, the Florida Keys are a hotspot for reef-based tourism , and local reefs are thus suering from a wide range of anthropogenic disturbances, including degraded water quality due to land-based sources of pollution , and high intensities of boating and diving activities . The Coral Restoration Foundation (CRF) was created in 2007 with the speciﬁc objective of growing and restoring threatened species of corals in the genus Acropora (A. cervicornis and A. palmata). Abundances of these two species of corals have declined by up to 90% throughout the Caribbean and both have been listed as “critically endangered” since 2008 by the IUCN . The Foundation harvests coral fragments from remnant colonies surviving on the reef and places them in coral nurseries. Early nursery prototypes included seaﬂoor table nurseries, but they then developed coral tree nurseries that are suspended in the water column at approximately eight meters depth (Figure 2C). Once fragments are large enough, they are planted directly onto the reef substrata using a 2-part marine epoxy cement (Figure 2C). Restoration eorts extend over 31 sites on 10 reefs along the upper Florida Keys reef tract  (Figure 1). 2.1.4. Box 4. Coral Restoration in St Croix, US Virgin Islands St Croix is a comparatively large high island (218 km in area) forming part of the US Virgin Islands in the Caribbean. Reefs around St Croix have suered extensively from climate change-related disturbances, similar to those described above for the Florida Keys. Tropical storms in 1989 and 1995 caused extensive reef damage, and several coral disease outbreaks over the past 20 years have caused further coral mortality [43,44]. In comparison to the Florida Keys, however, reefs around St Croix are not suering from intense tourism pressure. The Nature Conservancy (TNC) commenced coral restoration eorts in 2009, with the goal of growing and re-stocking endangered species of Acropora on local reefs . Initially, corals were collected as fragments of opportunity that had been broken from parent colonies naturally by storm or surge events. Currently, fragments are collected from donor colonies and grown in coral tree nurseries, Diversity 2020, 12, 153 6 of 22 following methods developed by CRF in Florida. Once fragments are large enough, they are planted back onto the reef using a 2-part marine epoxy cement. Restoration sites are scattered around the Island, with a particular focus on A. cervicornis restoration on the North Shore reefs of Cane Bay, and on A. palmata restoration near Green Cay and Knights Bay (Figures 1 and 2D). 2.2. Measuring ‘Snap Shots’ of Coral Assemblages’ Response to Restoration At each of the four locations, benthic data were compared among replicate restored sites (R), unrestored sites (UR), and control reference sites (CR). Sites were carefully selected with local reef managers to ensure we had the best representation for all three categories. At restored sites, coral fragments had been transplanted either directly onto the substrata or onto artiﬁcial structures. Unrestored sites were degraded sites directly adjacent to restored sites but were not the subject of coral restoration eorts. Unrestored sites were used as the direct control against which to assess potential eects of the restoration eort. Control reference sites were comparatively undisturbed sites nearby that were exposed to similar environmental conditions, thus their reef communities were hypothesized to be similar to those at the R and UR sites prior to degradation. CR sites acted as an indirect control providing some additional reference of the natural variability in reef condition at each particular location. A minimum of three replicate sites were surveyed for each of the three treatments (R, UR, CR) at each location, except at St Croix, where the extent of appropriate undisturbed reef area was so small that we could only survey two control reference sites. Thus, three restored sites, three unrestored sites, and three healthy reference sites were surveyed at all locations (except for the two CR sites at St Croix). In addition, a fourth restored site and a fourth unrestored site were surveyed in St Croix. Benthic data were recorded along three 20-m transect lines at each of the three sites per treatment in Koh Tao, Landaa Giraavaru, and the Florida Keys, for a total of 180 m surveyed per treatment at each of these locations. In St Croix, the restored area was too small for three replicate 20-m transects, thus two replicate 22.5-m transects were surveyed at each of four R and four UR sites (i.e., 180 m surveyed per treatment) to match the overall transect lengths surveyed at other locations. 2.2.1. Benthic Cover and Structural Complexity Benthic cover was measured using the line-intercept method, whereby the length of each substrate category falling directly under the line was recorded to the nearest centimeter . Substrate categories included all scleractinian corals, which were identiﬁed to the genus level; soft corals; macro-algae; other substrata like sand, rubble, and rocks; and other organisms, such as sponges, corallimorphs, and zoanthids. Percent cover of each substrate category was then calculated relative to the total length of each transect. Structural complexity of the reef under each transect line was scored qualitatively using a scale from 0 to 5, where 0 = no relief, and 5 = high structural complexity and high coral cover, following methods described in [20,47]. 2.2.2. Coral Health, Generic Richness, and Juvenile Recruitment In addition to line-intercept surveys of coral cover, 2-m-wide belts were surveyed along each transect line (i.e., a 40-m area per transect), within which all hard corals were identiﬁed to the genus level and assigned to a coral health category. Corals were scored as either healthy or having signs of one or more disease types, and/or a range of compromised health states, such as algal overgrowth, sediment smothering, physical damage, or signs of predation. The prevalence of each disease or compromised health category was calculated as its percentage relative to the total number of coral colonies surveyed in each 40-m belt transect. Coral health categories and assessment protocols followed guidelines developed for the Indo-Paciﬁc , and for Caribbean reefs . These survey techniques have been applied previously to assess coral health (e.g., [31,50]). The number of coral genera recorded in each belt transect was used as a measure of generic richness. The number of coral Diversity 2020, 12, 153 7 of 22 juveniles (colonies with a diameter under 5 cm; ) was also recorded within each belt transect, and used as a proxy for the number of coral recruits in recent years . 2.3. Data Analysis All data were analyzed using the statistical program R (version 3.4.1, ). The analyses described below were applied to metrics measured at each of the four locations separately. Given the large geographic distances between the four locations and inherent dierences in biodiversity and coral cover among their reef communities, only qualitative comparisons of the summative results are made among the four reef locations. 2.3.1. Benthic Cover For each of the four locations, the mean percent cover of each substrate category was compared among treatments (R, UR, and CR) and sites (n = 3 or 4 sites per treatment type) using multi-factor general linear models (GLMs). Treatments were analyzed as ﬁxed factors and sites as random factors. A variety of models were tested, including ones where explanatory variables were treated as having either additive or multiplicative eects, and where data were log-transformed. AICc model selection was used to select the model explaining the greatest variation in the data, i.e., the model having the lowest AICc score. Assumptions for model validity were checked through Q-Q plots and residual plots. When tests failed to meet the assumptions of a Gaussian distribution after log-transformation, non-parametric Kruskal–Wallis tests were applied. When applicable, post-hoc Tukey’s HSD tests were also applied to tease out dierences among treatments and sites. 2.3.2. Structural Complexity Analyses of the mean structural complexity scores among treatments and sites at each location were performed using multi-factor general linear models, as described above for benthic cover analyses. 2.3.3. Coral Generic Richness Abundance of Juvenile Corals Multi-factor general linear models were also used to compare generic richness and abundance of juvenile corals among treatments and sites at each location. Details of analyses and checks of assumptions were as described above for benthic cover data, except that data were modelled as having “Poisson” or “negative binomial” distributions, as these are the most appropriate distributions for count data. Analysis of coral juvenile abundance could only be done for two of the four sites: Koh Tao and Landaa Giraavaru, as sites in the Florida Keys and St Croix were data deﬁcient for this indicator, i.e., there was little to no recruitment at any of these sites. 2.3.4. Coral Health Percentages of corals in each health category were compared among treatments and sites using analyses similar to those described above for benthic cover. Prevalence values for each of four health categories were compared among treatments and sites at each location, namely the prevalence of healthy corals, diseased corals, corals with signs of predation, and corals with other signs of compromised health. 2.3.5. Coral Assemblages Multivariate analyses were used to assess potential dierences in the composition of coral assemblages (i.e., abundance of local coral genera and other benthic categories) among treatments at each location. Prior to analysis, all data were transformed using Wisconsin’s double transformation for the fourth root. We then created distance matrices based on “Bray-Curtis” dissimilarity indices, as these are good at detecting ecological gradients , and applied non-metric multidimensional scaling (nMDS) to the transformed dataset. The validity of the nMDS was checked through evaluation Diversity 2020, 12, 153 8 of 22 of the R value of the linear and non-linear ﬁt, as well as the stress value, which was assumed to be good when <0.2 . Coral health and benthic cover data were overlaid on top of the nMDS, and ADONIS tests (multivariate ANOVA based on dissimilarities) were used to calculate the contribution of each variable to the spread of the benthic community data, as well as to the dierences in the composition of coral assemblages between sites at each location (pairwise ADONIS). Finally, SIMPER analyses were performed to reveal the cumulative contributions of the most inﬂuential coral genera and benthic category to the spread of the data at each location. 3. Results 3.1. Hard Coral Cover Mean hard coral cover was more than twice as great at restored treatments compared to degraded unrestored treatments at three of the four locations: Koh Tao (LM: F = 9.5 p < 0.001, Table S1), Landaa Giraavaru (LM: F = 6.9, p < 0.001, Table S1), and the Florida Keys (GLM: Residual Deviance = 9.4, p = 0.005, Table S1, Figure 3). In St Croix, there was a trend towards higher hard coral cover at restored treatments compared to unrestored treatments, but the dierence was not statistically signiﬁcant (GLM: RD = 1.7, p = 0.375, Table S1, Figure 3). Figure 3. Mean percent cover of hard corals per 40-m belt transect (SE) compared among treatments (unrestored, restored, control reference sites) at each of the four locations. Letters above each histogram indicate whether mean values dier signiﬁcantly (dierent letters) or are statistically indistinguishable (same letters). n = 9 transects per treatment in Koh Tao, Landaa Giraavaru, and the Florida Keys; In St Croix, n = 8 transects for unrestored and restored treatments, n = 6 transects for the control reference treatment. There were trends for absolute values of mean hard coral cover to be higher at restored treatments than at control reference treatments at the two Indo-Paciﬁc locations (Koh Tao and Landaa Giraavaru); conversely, means were highest at control reference treatments at both Caribbean locations (Florida and St Croix; Figure 3). However, at all four locations, dierences in mean hard coral cover between restored treatments and control reference treatments were not statistically signiﬁcant (Figure 3, Table S1). Diversity 2020, 12, 153 9 of 22 3.2. Structural Complexity Structural complexity of the coral community was signiﬁcantly higher at restored treatments compared to unrestored degraded treatments at all four locations (Figure 4, Table S2). In Koh Tao, structural complexity scores were 2 times greater at restored compared to unrestored treatments (LM: F = 23.18, p < 0.001, Table S2), and 1.5 times greater at restored compared to control reference treatments (GLM: p = 0.0013, Table S2). At all three other locations, although structural complexity scores were 1.5 times greater at restored than at unrestored treatments (Landaa Giraavaru LM: F = 6.9, p = 0.0014, Florida Keys LM, F = 11.5, p = 0.019, St Croix LM, F = 19.4, p < 0.001, Table S2), and mean scores were highest at control reference treatments (although not signiﬁcantly dierent; Figure 4). Scores at restored treatments were consistently greater than the average score across all sites (2.5 out of 5), whereas scores at unrestored treatments were consistently lower (Figure 4). Figure 4. Mean structural complexity scores (SE) compared among treatments (unrestored, restored, control reference sites) at each of the four locations. Letters above each histogram indicate whether mean values dier signiﬁcantly (dierent letters) or are statistically indistinguishable (same letters). n = 9 transects per treatment in Koh Tao, Landaa Giraavaru, and the Florida Keys; in St Croix, n = 8 transects for unrestored and restored treatments, n = 6 transects for the control reference treatment. 3.3. Number of Coral Juveniles This indicator was only valid for Koh Tao and Landaa Giraavaru because juvenile coral colonies were not detected in high enough abundance in the Florida Keys or St Croix to provide sucient data for statistical analyses. In Koh Tao, the mean abundance of juvenile corals was greatest at restored treatments. Although abundances at restored treatments were not signiﬁcantly greater than those at control reference treatments (Table S3, Figure 5), they were signiﬁcantly greater than those at unrestored treatments, where no juveniles were recorded (Kruskal–Wallis: = 8.22, df = 2, p = 0.043, Table S3, Figure 5). Overall, the mean number of juveniles recorded in Koh Tao was 5.7/40 m , with abundances diering among restored treatments according to the artiﬁcial structures used (Kruskal–Wallis: = 6.06, Diversity 2020, 12, 153 10 of 22 df = 2, p = 0.049, Table S4). The highest number of juveniles recorded were on concrete reef balls in Tanote Bay (Figure 6), and the lowest number recruited to the mix of steel frames and bottle nurseries in Chalok Bay (Figure 6). In Landaa Giraavaru, the mean abundance of coral juveniles (8 juveniles/40 m across all sites and treatments) did not signiﬁcantly dier among the three treatments (Kruskal–Wallis: = 0.825, df = 2, p = 0.66; Figure 5, Table S3). Figure 5. Mean number of juvenile corals counted per 40-m belt transect (SE) compared among treatments (unrestored, restored, control reference sites) in Koh Tao (Thailand) and Landaa Giraavaru (Maldives). Letters above each histogram indicate whether mean values dier signiﬁcantly (dierent letters) or are statistically indistinguishable (same letters). n = 9 transects per treatment. Figure 6. Mean number of juvenile corals counted per 40-m belt transect (SE) compared among the three restored sites in Koh Tao (Thailand). Restoration designs varied among the three sites, i.e., corals were transplanted: onto electriﬁed steel frames at the Biorock site, onto steel frames and into glass bottles embedded in concrete in Chalok, and onto concrete reef balls in Tanote. Letters above each histogram indicate whether mean values dier signiﬁcantly (dierent letters) or are statistically indistinguishable (same letters). n = 9 transects per treatment. 3.4. Coral Generic Richness Increases in coral generic richness at restored compared to unrestored treatments only occurred in Koh Tao (GLM: RD = 13.2, p = 0.0352, Table S5, Figure 7). In both the Florida Keys and St Croix, Diversity 2020, 12, 153 11 of 22 coral generic richness was similar across all treatments at all locations (Table S5, Figure 7). In Landaa Giraavaru, coral generic richness was signiﬁcantly lower at the restored treatments compared to both unrestored (GLM: RD = 29.2, p = 0.0015) and control reference treatments (GLM: RD = 29.2, p < 0.001), (Table S5, Figure 7). Figure 7. Mean number of coral genera per 40-m belt transect (SE) compared among treatments (unrestored, restored, control reference sites) at each of the four locations. Letters above each histogram indicate whether mean values dier signiﬁcantly (dierent letters) or are statistically indistinguishable (same letters). n = 9 transects per treatment in Koh Tao, Landaa Giraavaru, and the Florida Keys; in St Croix, n = 8 transects for unrestored and restored treatments, n = 6 transects for the control reference treatment. 3.5. Coral Health Coral health varied among treatments and locations. In Koh Tao, unrestored treatments had a four-fold higher prevalence of unhealthy coral colonies compared to both restored and control reference treatments (GLM: RD = 1534, p < 0.001, Table S6), which was driven by a four-fold higher prevalence of coral colonies with signs of compromised health (GLM: RD = 4.35, p < 0.001, Table S8, Figure 8). The prevalence of diseased corals and of colonies with signs of predation did not dier among treatments (Figure 8, Table S7). Signs of predation in Koh Tao were primarily identiﬁed as feeding scars from Drupella snails and crown-of-thorns starﬁsh (COTS). In Landaa Giraavaru, the prevalence of unhealthy coral colonies was consistently over 80% of all colonies in all treatments. The overall high prevalence of unhealthy corals was driven by a high (62.4%) mean prevalence of bleached corals. Disease prevalence was also two times greater at restored compared to control reference treatments (GLM: RD = 6.03, p = 0.025, Figure 8, Table S7). In the Florida Keys, disease prevalence was highest at control reference treatments; disease prevalence was 1.5 times greater than at restored sites (GLM: RD = 1.64, p = 0.028, Table S7), and 2.8 times greater than at unrestored treatments (GLM: RD = 1.64, p = 0.006, Table S7, Figure 8). Only restored treatments had signs of predation, thus the prevalence of predation scars was signiﬁcantly higher at these treatments compared to both unrestored (Kruskal–Wallis, Chi-square = 21.034, df = 2, Diversity 2020, 12, 153 12 of 22 p = 0.038, Table S9) and control reference treatments (Kruskal–Wallis, Chi-square = 21.034, df = 2, p = 0.038, Table S9, Figure 8). In St Croix, restored treatments had a higher prevalence of diseased colonies than unrestored (GLM: RD = 0.41, p < 0.001, Table S7) and control reference treatments (GLM: RD = 0.41, p = 0.037, Table S7), and a higher prevalence of colonies with other signs of compromised health than control reference treatments (GLM, RD = 0.92, p < 0.001, Table S8, Figure 8). Restored treatments were also the only sites at which we observed signs of predation (Figure 8). Signs of predation in both the Florida Keys and St Croix were dominated by scars from ﬂatworms and ﬁsh bites. Figure 8. Mean prevalence of corals in four health categories representing unhealthy states (corals with signs of disease, bleaching, predation, or other signs of compromised health) per 40-m belt transect compared among treatments (unrestored, restored, control reference sites) at each of the four locations. n = 9 transects per treatment in Koh Tao, Landaa Giraavaru, and the Florida Keys; In St Croix, n = 8 transects for unrestored and restored treatments, n = 6 transects for the control reference treatment. 3.6. Composition of the Coral Assemblages The composition of coral assemblages diered among restoration treatments at all four locations. In Koh Tao, the composition of coral assemblages at control reference treatments diered signiﬁcantly from those at both restored and unrestored treatments (ADONIS: (CR to R) F = 3.64, p = 0.014; (CR to UR) F = 4.52, p = 0.008, Table S10, Figure 9). There was also a signiﬁcant eect of sites on the composition of the coral assemblages (ADONIS: F = 5.67, p = 0.001). Overall, coral assemblage composition at the restored treatments was intermediate between those at the unrestored and control reference treatments (Figure 9). Restored treatments had four times more cover of corals in the family Acroporidae than either the unrestored or control reference treatments (Figure 10). Accordingly, the cumulative contribution of Acroporidae accounted for 75% of the dierences between restored and unrestored treatments (SIMPER). Sand dominated the benthos at unrestored treatments, accounting for 47% (SIMPER, cumulative contributions) of the dierences between unrestored and restored treatments, and 38% of the dierences between unrestored and control reference treatments (SIMPER, cumulative contributions). Poritidae and Fungiidae were also more abundant at control reference treatments than restored and unrestored treatments (Figure 10). Diversity 2020, 12, 153 13 of 22 Figure 9. Eect of coral restoration on the composition of coral assemblages at four geographic locations, as represented by non-metric multidimensional scaling. Polygons represent coral assemblages in each treatment, where green polygons encompass restored sites, blue polygons encompass unrestored sites, and grey polygons encompass control reference sites. Colored shading reﬂects the location of the respective set of sites in non-metric multi-dimensional scaling space. Vectors represent the inﬂuence of benthic attributes on the benthic community composition. Figure 10. Comparisons of the mean cover of the most inﬂuential substrate categories (post-simper analyses) per 40-m belt transect among treatments (unrestored, restored, reference control sites) at each of the four locations. n = 9 transects per treatment in Koh Tao, Landaa Giraavaru, and the Florida Keys; In St Croix, n = 8 transects for unrestored and restored treatments, n = 6 transects for the control reference treatment. Diversity 2020, 12, 153 14 of 22 In Landaa Giraavaru, the composition of coral assemblages at the restored treatments diered signiﬁcantly from the composition of assemblages at unrestored and control treatments (ADONIS: (R to UR) F = 3.33, p = 0.15; (R to CR) F = 3.78, p = 0.005, Table S10). Coral assemblages were also signiﬁcantly dierent at control compared to unrestored treatments (ADONIS: F = 2.29, p = 0.045, Table S10). There was also a signiﬁcant eect of site on the composition of coral assemblages (ADONIS: F = 2.18, p = 0.004). Restored treatments were characterized by higher cover of corals in the family Acroporidae and by lower cover of rubble (Figure 10). Rubble contributed to 30% of the dierences between restored and unrestored treatments, and 72% of the dierences between restored and control reference treatments (SIMPER, cumulative contributions). Acroporidae contributed 58% of the dierences between restored and unrestored treatments, and 55% of the dierences between restored and control reference treatments (SIMPER, cumulative contributions). In the Florida Keys, only unrestored treatments had a distinct benthic community composition (ADONIS: (UR to R) F = 3.52, p = 0.014; (UR to CR) F = 3.88, p = 0.006, Table S10, Figure 9). There was also a signiﬁcant site eect on the composition of the benthic community (ADONIS: F = 3.88, p = 0.001). In terms of benthic composition, rocks, gorgonians, and Acroporidae were the most inﬂuential factors driving dierences among treatments (SIMPER). The cover of corals in the family Acroporidae was nil at unrestored treatments, and highest at control reference treatments. Acroporidae accounted for 56% of the dierences between unrestored and control treatments, 84% of the dierences between unrestored and restored treatments (SIMPER, cumulative contribution), and 64% between restored and control reference treatments (SIMPER, cumulative contribution) (Figure 10). Rocky substrate and gorgonians had the highest percent cover in unrestored treatments (Figure 10). Rocky substrate accounted for 32% of the dierences between unrestored and restored treatments, and 80% of the dierences between unrestored and control reference treatments (SIMPER, cumulative contribution). Gorgonian cover was twice as high in unrestored compared to both restored and control reference treatments and thus accounted for 65% of the dierences between unrestored and restored treatments, and 29% of the dierences between unrestored and control reference treatments (SIMPER, cumulative contribution) (Figure 10). In St Croix, the coral assemblages at restored treatments diered signiﬁcantly from those of both unrestored and control reference treatments (ADONIS: (R to UR) F = 6.96, p = 0.001; (R to CR) F = 3.5, p = 0.004, Table S10). The coral assemblages at control reference treatments were also distinct from those of the unrestored treatments (ADONIS: F = 3.15, p = 0.017, Table S10). Benthic community composition also varied signiﬁcantly among sites (ADONIS: F = 3.49, p = 0.001). Dierences in benthic community composition were driven by the cover of corals in the family Acroporidae, which was 1.9 times greater at restored than at control reference treatments; Acroporidae corals were absent at unrestored treatments (Figure 10). Acroporidae therefore accounted for 71% of the dierences between unrestored and restored treatments (SIMPER, cumulative contribution). Astrocoeniidae were only present in control reference treatments and accounted for respectively 69% and 65% of the dierences in the benthic community between restored and control reference treatments, and between unrestored and control reference treatments (SIMPER, cumulative contribution). Benthic communities at unrestored treatments were dominated by the presence of rocks and algae (Figure 10). The eects of benthic attributes on the composition of coral assemblages at all four locations is further explored in the Supplementary Material (Section S2). 3.7. Summary and Links with Restoration Designs The eect of coral restoration on the ﬁve ecological indicators surveyed diered among the four study locations, reﬂecting dierences in restoration designs and local factors (Figure 11). While our snapshot surveys prevent us from linking restoration outcomes to speciﬁc designs, some designs may have warranted stronger responses than others. For example, all ﬁve indicators surveyed positively increased in restored treatments in Koh Tao, where the restoration design includes a mix of direct transplantation and a variety of artiﬁcial structures (steel frames, electriﬁed steel frames, concrete reef Diversity 2020, 12, 153 15 of 22 balls, and glass bottles in concrete). This combination of techniques led to the highest rate of increase in structural complexity, coral generic diversity, number of juveniles, and improved coral health at restored compared to unrestored treatments of all study locations (Figure 11). Other designs that used artiﬁcial structures like the steel frames in Landaa Giraavaru also led to signiﬁcant increases in hard coral cover and structural complexity at restored compared to unrestored treatments (Figure 11), but restoration outcomes at this location also included signiﬁcant decreases in coral generic richness at restored treatments. Direct transplantation was the only technique used in both the Florida Keys and St Croix. This technique resulted in consistent increases in hard coral cover, structural complexity, and coral generic diversity (Figure 11). In the Florida Keys, the restoration design also led to ﬁve times greater hard coral cover at restored compared to unrestored treatments, thus this metric increased by the greatest amount in the Florida Keys out of all four study locations (Figure 11). Conversely, increases in hard coral cover at restored compared to unrestored treatments were the lowest in St Croix (Figure 11). Finally, coral health was poorer in restored compared to unrestored treatments in both the Florida Keys and St Croix. Figure 11. Summary table comparing ﬁve ecological indicators surveyed at four study locations with dierent restoration designs. Numerical values represent the ratios of each metric at restored compared to unrestored treatments. Colored boxes represent the signiﬁcance of the dierence between restored and unrestored treatments. Green denotes signiﬁcant positive ratios; red denotes signiﬁcant negative ratios; blue denotes non-signiﬁcant dierences. 4. Discussion This study is the ﬁrst to examine the long-term eects of coral restoration practices on coral assemblages and to test the generality of outcomes across programs using diering protocols in a range of geographic locations. We found systematic increases in hard coral cover and reef structural complexity at restored compared to unrestored treatments at all four locations surveyed. Moreover, multivariate analyses conﬁrmed that outplanted corals had substantial impacts on local benthic communities, resulting in the community composition at restored sites diering from that of unrestored Diversity 2020, 12, 153 16 of 22 and comparatively healthy control sites. Patterns in the responses of other ecological indicators of reef resilience to restoration varied across locations, potentially reﬂecting variations in local benthic assemblages and/or variations in community responses to dierent restoration methodologies. 4.1. Restoration Increases Coral Cover and Structural Complexity The doubling of hard coral cover at restored compared to unrestored treatments at three out of four locations indicates that the range of restoration techniques investigated here are eective strategies for restoring coral assemblages. Moreover, coral cover was higher in restored plots than at control reference treatments following 10 years of restoration at both Indo-Paciﬁc locations (Koh Tao and Landaa Giraavaru). While coral cover remained highest at control reference treatments in the Florida Keys and St Croix, the restoration goals of these two Caribbean programs were more focused on growing and restoring endangered species of Acropora (A. cervicornis and A. palmata) . Systematic increases in hard coral cover at restored treatments are unsurprising, as corals fragments were actively planted at all four locations. However, results suggest that while corals may suer post-transplant stress and mortality [56–58], restoration eorts at all four locations are substantive enough to have positive eects on coral cover over 10-year timeframes. Increased hard coral cover is a necessary ﬁrst step towards increasing reef resilience, increasing local breeding populations of corals, providing habitats for juvenile ﬁsh and invertebrates, and potentially preventing or at least mitigating phase shifts towards algae-dominated systems [1,59]. Signiﬁcant increases in structural complexity at restored compared to unrestored treatments at all four study locations suggest that both direct transplantation of coral fragments onto the reef substrata and transplantation onto artiﬁcial structures are eective in increasing reef complexity at restored sites. In Koh Tao, where coral fragments are generally attached to artiﬁcial structures, structural complexity was doubled at restored compared to unrestored treatments, and higher at restored compared to control reference treatments. Although artiﬁcial structures were used in Landaa Giraavaru, structural complexity did not dier signiﬁcantly between restored and reference treatments, largely because of the high natural complexity of control reference reefs (mean structural complexity greater than 4 out of 5). Here, complexity represents the degree of reef relief (cf. ) but does not speciﬁcally account for the number and sizes of holes and crevices present in the reef matrix, which may aect the abundance and diversity of ﬁsh and invertebrates . In the Florida Keys and St Croix, a lack of dierence in structural complexity between restored and control reference treatments reﬂects the fact that most of the complexity at these locations is provided by the presence or absence of thickets of branching Acropora, which are the targets of the restoration eorts . 4.2. Other Coral-Based Indicators of Reef Resilience Vary among Restoration Programs Despite increases in coral cover and structural complexity at restored treatments, other critical indicators of reef resilience did not increase consistently in response to restoration eorts. For example, higher densities of juvenile corals at restored compared to unrestored treatments were only found in Koh Tao, and only on concrete reef balls. It may be that the high surface rugosity of reef balls is conducive to coral larval settlement [62,63]. However, because Koh Tao was the only restoration program out of the four studied to use these structures, and they were only used at one out of the three restored sites, we are unable to distinguish between the potential contributions of site versus type of structure on the increased abundance of coral juveniles at this one site. In Landaa Giraavaru, lack of dierences in juvenile coral density among treatments might be attributable to either the type of structure used (i.e., steel frame structures that are not conducive to larval settlement), and/or the fact that reefs around the island are limited by recruitment. Here, the average number of juveniles recorded across all sites (0.8/m ) was much lower than coral recruit densities previously reported in the Maldives (2.5 to 18 recruits/m , ), and in other regions of the world (4 to 80 recruits/m , [64,65]). However, these studies deﬁne coral recruits as any new corals colonizing the restored sites , and use dierent survey techniques (e.g., recruitment tiles, ). It is possible that Diversity 2020, 12, 153 17 of 22 the methods used here, of only recording corals with a diameter <5 cm in 2-m belt transects, may have limited the detection of coral recruits. This interpretation is supported by ﬁndings of similar densities of recruits on Lord Howe Island reefs using the same methods . The paucity of recruitment in both the Florida Keys and St Croix precluded investigating the eect of coral restoration on the abundance of juvenile corals at these two locations, and further conﬁrms that reefs in the Caribbean are severely limited in their ability to recruit new juvenile corals [66,67]. Coral generic richness was a second indicator of reef resilience that was not consistently augmented by restoration programs. Coral restoration only positively aected coral generic richness in Koh Tao, where the restoration design explicitly aims to maximize the diversity of coral transplants. In the three other locations, targeted transplantation of speciﬁc corals meant that coral generic diversity was either lowest at the restored treatments (Landaa Giraavaru) or indistinguishable from unrestored treatments (Florida Keys, St Croix). In Landaa Giraavaru, coral transplants were dominated by fast-growing branching corals from the genera Acropora and Pocillopora, artiﬁcially boosting the density of these two genera at restored sites. The lack ofa restoration eect on the generic richness in the Florida Keys and St Croix was unsurprising given that restoration eorts target the two endangered species of Acropora . Finally, coral health, a third indicator of reef resilience that was not consistently improved by restoration, revealed location-speciﬁc patterns. Again, this indicator was improved only in Koh Tao, potentially because of the high level of maintenance of the restoration sites by the NHRCP team. It is also likely that elevation of the corals slightly above the substrata on artiﬁcial structures prevented them from being smothered by sediments or algae. Unrestored treatments had a signiﬁcantly higher prevalence of colonies with sediment damage and algal overgrowth (included in the category ‘other signs of compromised health’), corroborating this line of reasoning. It is noteworthy that there was no evidence that transplanted fragments were more susceptible to disease due to manipulation and injuries sustained in the process of attaching them to structures. In summary, results from Koh Tao suggest that planting corals above the substrata and maximizing the diversity of corals transplanted are good strategies to maximize coral health at restored sites. In Landaa Giraavaru, poor coral health in all treatments reﬂected that, at the time of the survey, the Maldives were experiencing mass coral bleaching. Corals at all survey locations were severely bleached regardless of the depth or restoration treatment. The overriding impact of thermal stress at the time of the surveys is a reminder that active intervention approaches like coral restoration cannot prevent climate-driven exposure events that overwhelm reef resilience. In the Florida Keys, coral disease prevalence was highest at control reference treatments, potentially because of high densities of Acropora combined with no active maintenance of natural reef areas, and the overall history of disease-related loss of Caribbean species of Acropora [39,68]. The prevalence of predation scars, on the other hand, was highest at the restored treatments, likely reﬂecting ﬁreworm predation on freshly planted A. cervicornis [61,69,70]. In St Croix, restored treatments were again the only sites to experience coral predation at that location. Together with higher disease prevalence, restored treatments had overall lower coral health than either unrestored or control reference treatments. Results from both the Florida Keys and St Croix raise questions about whether Acroporids are good candidates for coral restoration in the Caribbean. While the two Caribbean programs are meeting their goal of increasing Acropora cover at restored sites , focusing on this genus alone might not lead to successful long-term outcomes in terms of reef resilience and enhanced reef-related ecosystem goods and services. Maximizing the diversity of coral transplants at these locations might help harness natural ecological processes that decrease competition between and predation upon freshly transplanted corals, and therefore optimize the long-term outcomes of the restoration process [21,72]. Diversity 2020, 12, 153 18 of 22 4.3. Coral Restoration Inﬂuences the Composition of the Benthic Community Restoration aected the composition of benthic communities at all four locations, highlighting that coral restoration eorts can aect communities at a much greater scale than that of the coral transplants. This result supports the idea that characterizing restoration eectiveness requires broad reef-scale considerations . Restoration methodologies, from site selection to the use of artiﬁcial structures and the species and density of coral transplants used, all require careful consideration in terms of their impact on local benthic communities. Site selection, in particular, is increasingly recognized as an important factor for maximizing the outcomes of restoration eorts [10,21,72]. Comparison of the benthic community composition between restored and control reference sites is a useful indicator of the appropriateness of the site selected. For example, similarities in benthic community assemblages between restored and control reference sites in the Florida Keys suggest that restoration eorts increased the resilience of benthic communities at these sites, and that site selection for the restoration eort was indeed appropriate, even given the degree of natural degradation at the control reference sites. 5. Limitations and Further Research This study represents a “snapshot” of the responses of coral assemblages to restoration practices and our data and remarks on reef resilience should be interpreted within this context. Because of our sampling design comparisons of restoration eectiveness among the four programs are limited due to the variable restoration designs, the level of transplant maintenance, and the age of restored plots all varied among the four locations. Likewise, in three of the programs, only one type of restoration design was used (i.e., metal frames in the Maldives, midwater nurseries at both Caribbean locations), precluding meaningful comparisons of restoration eectiveness between designs. There is scope for small-scale research at a particular location on local indices of restoration eectiveness among dierent types of artiﬁcial structures and between artiﬁcial structures versus direct transplantation onto reef substrata at one location to complement our broad geographic comparisons. Furthermore, data were collected at the genus rather than species level so that restoration managers could easily replicate our monitoring program. However, species-level data, especially when quantifying coral juveniles in terms of success/recruitment, would provide greater insights into changes in coral diversity patterns and impacts on coral health, especially for restoration programs focused on restoring endangered coral species (e.g., Acropora species in the Caribbean). Overall, our research reveals that planting corals onto degraded reefs results in consistent long-term increases in hard coral cover and reef structural complexity, both of which are necessary steps in the recovery of degraded reefs, a major goal of restoration programs. The results presented here thus demonstrate that the potential for coral restoration eorts to increase coral reef resilience in the long-term is promising, but restoration practices should focus more closely on maximizing coral generic richness, as well as planting corals o the substrata or in low-predation areas to maximize coral health at restored sites. The eectiveness of coral restoration eorts also needs to account for its eects on other important functional groups, such as ﬁshes , and further explore factors that contribute to enhancing the generic diversity, fecundity, and recruitment of juvenile corals at the restored sites. Considerations of socio-economic factors will also be critical in assessing the potential of coral restoration to contribute to resilience-based reef management [4,5]. Diversity 2020, 12, 153 19 of 22 Supplementary Materials: The following are available online at http://www.mdpi.com/1424-2818/12/4/153/s1, Table S1. Hard coral cover among treatments Posthoc with Tukeys’ contrast on linear models, Table S2. Structural complexity among treatments. Posthoc with Tukeys’ contrast on linear models, Table S3. Coral juveniles among treatment. Posthoc on Kruskal Wallis with Nemenyi test, Table S4. Coral juveniles among restored sites in Koh Tao. Posthoc on Kruskal Wallis with Nemenyi test, Table S5. Coral generic richness among treatments. Posthoc with Tukeys’ contrast on general linear models and Kruskal Wallis with Nemenyi test, Table S6. Coral health prevalence among treatments. Posthoc with Tukeys’ contrast on general linear models, Table S7. Coral disease prevalence among treatments. Posthoc with Tukeys’ contrast on general linear models, Table S8. Prevalence of compromised coral colonies among treatments. Posthoc with Tukeys’ contrast on general linear models, Table S9. Prevalence of predated upon coral colonies among treatments. Posthoc with Tukeys’ contrast on general linear models, and Kruskal Wallis Nemenyi tests, Table S10. Pairwise ADONIS investigating the compositional dierences in coral assemblages among restoration treatments at the four program locations calculated from Bray-Curtis distance matrices, Section S2. Eects of benthic attributes on the compositional dierences of coral assemblages. Author Contributions: Study design, M.Y.H., R.B., A.B., N.M.G., N.M., B.L.W.; Data collection, M.Y.H., T.L.B., J.L., C.M.S., L.T.; On-site logistics for site selection and data collection, J.L., T.L.B., C.M.S., L.T.; manuscript writing, M.Y.H., R.B., A.B., N.M.G., N.M., B.L.W.; manuscript review and editing, all authors. All authors have read and agreed to the published version of the manuscript. Funding: This research was supported by the College of Science and Engineering at James Cook University and the ARC Centre of Excellence for Coral Reef Studies. In the Florida Keys, data were collected following requirements of the Coral Restoration Foundation’s research policies (ID CRF-2016-23), and CRF’s permit FKNMS-2011-159-A4. Acknowledgments: We would like to thank P. Urgell, K. Magson, F. Couture, S. Stradal, K. Ripple, R. Willis, K. Lewis, K. Kopecki, and J. Blomberg for assistance in data collection, as well as T.J. Chase for useful comments and improvements to the manuscript. 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