Coral reefs, often characterized as underwater rainforests, are among the most biodiverse ecosystems on Earth (Reaka-Kudla, 1997; Jaap, 2000) and perform critical ecological and economic functions for millions worldwide (e.g., Moberg and Folke, 1999; Ferrario et al., 2014). However, dramatic declines in coral cover globally over the past several decades (Gardner et al., 2003; Côté et al., 2005; De’ath et al., 2012) have eroded the functionality of these ecosystems and gravely threaten them and the services they provide (Jones et al., 2004; Alvarez-Filip et al., 2011). These declines have been driven by both natural and anthropogenic stressors such as overfishing, disease, storms, nutrient pollution, coastal development, and climate change (e.g., Gladfelter, 1982; Hughes, 1994; Bruno et al., 2007; Burke et al., 2011; Walton et al., 2018). Reefs will continue to decline as stressors such as bleaching, ocean acidification, and overall impacts of climate change become more prevalent (Hoegh-Guldberg et al., 2007; Albright et al., 2010; Hughes et al., 2017).

The Florida Reef Tract (FRT), located at the southern terminus of the Florida Peninsula spanning from the Florida Keys up the eastern part of the state to Martin County, is the third largest barrier coral reef on Earth and the only barrier coral reef in the continental United States. As of 2015, the Florida Keys National Marine Sanctuary located within the FRT produced an estimated $4.2 billion in economic output to the state of Florida and attracts over three million visitors each year (Tbd Economics, LLC, 2019). However, like most Caribbean reef systems, the FRT has suffered dramatic declines in coral cover from an estimated 25% in the 1980s to an average stony coral cover of 6% today (Ruzicka et al., 2013). Among the several dozen coral species that have dominated shallow water reefs in the western Atlantic for hundreds of thousands of years, elkhorn coral, Acropora palmata, was historically among the most abundant (Jackson, 1992; Precht and Miller, 2007). However, like other coral species, A. palmata has declined dramatically in Florida over recent decades. This decline, driven largely by outbreaks of white-band disease in the 1970s (Gladfelter, 1982; Aronson and Precht, 2001) and more recently by additional stressors including ocean warming, human activities, and storm damage (Baums et al., 2003; Burke et al., 2011; Williams and Miller, 2012; Williams et al., 2017), led to A. palmata being listed as critically threatened by the IUCN in 2008 (Aronson et al., 2008). No significant signs of recovery have been recorded (Miller et al., 2002; Johnson et al., 2011), and recent data suggest that A. palmata populations are now too sparsely distributed across the reef tract to successfully reproduce without direct intervention (Knowlton, 2001; Williams et al., 2008; Miller et al., 2018; National Academies of Sciences Engineering Medicine, 2019). More broadly, because no other Caribbean corals closely resemble A. palmata morphologically or ecologically, the only way in which Caribbean coral reefs could regain their architectural composition and structural and biological function would be to restore A. palmata specifically (Chamberland et al., 2015). Because species such as A. palmata have experienced relatively rapid declines over the last several decades (Bruckner, 2002; Miller et al., 2002; Sutherland and Ritchie, 2004), it is important for practitioners to grow corals quickly and efficiently to reach goals of population enhancement.

The dire state of coral reefs worldwide has led to a surge in active coral restoration efforts. These efforts, which have become increasingly effective (Young et al., 2012; Carne et al., 2016), provide a critical capability to help coral reefs survive the Anthropocene until the drivers of coral loss can be directly addressed. Generally, coral restoration efforts are classified either as sexual restoration, where individual genets are fertilized, grown, and released (or “outplanted”) into the wild, or asexual restoration, in which larger colonies are broken into clonal sub-fragments. Though hobbyists have been asexually fragmenting aquarium corals for many decades, the asexual restoration approach known as “coral gardening” has only been in use for a quarter century (Rinkevich, 1995). This approach, which involves raising corals in nurseries before ‘outplanting’ them back onto reefs, is practiced worldwide (e.g., Montoya-Maya et al., 2016), and includes over 150 programs in more than 20 countries in just the Caribbean and Western Atlantic alone (Lirman and Schopmeyer, 2016). Coral gardening has reached ecologically meaningful scales with 10,000s of coral grown in nurseries and outplanted each year (Lirman and Schopmeyer, 2016). To date, most restoration practitioners focus on asexual coral production through fragmentation (fragment size of several to many cm²) or microfragmentation (fragment size of ∼1 cm²) in part because it is relatively straight-forward in methodology, inexpensive, and can produce large amounts of biomass within short periods of time thus increasing the capability of restoring reefs at large scales. Another advantage of asexual restoration is the capability to propagate “winning” genets at large scales, maximizing the probability of restoration success and reef persistence.

Despite the large number of coral restoration practitioners now operating worldwide, and the environmental and social benefits of sharing best practices, and specific coral propagation methods are often poorly documented. Indeed, such methods often remain unpublished, or are relegated to hobbyist forums or gray literature (Barton et al., 2017), hampering effective, evidence based coral restoration (Boström-Einarsson et al., 2020). Accordingly, the publishing of best practices, failures, benchmarks, and even raw restoration data has been identified as a critical need to help develop the rapidly evolving field of coral restoration (Lirman and Schopmeyer, 2016; Boström-Einarsson et al., 2020).

There are many questions that remain regarding the best practices of coral restoration; among those are identifying how genet and substrate type alter the growth and survival of microfragments. Identifying superior growth substrates is critical for scaling up production at restoration facilities while keeping costs low. Similarly, understanding the various strengths and weaknesses of genets (including growth potential, resilience to disturbance, and reproductive potential) is particularly critical for asexual propagation programs, which often propagate fragments from a small number of genets in their nurseries. As new genets are obtained, practitioners may need to evaluate the characteristics of these genets (e.g., growth rate, stress tolerance) against those in the collection to prioritize propagation and outplanting. While genetic difference in growth across other species has been documented (Osinga et al., 2011; Lirman et al., 2014; Drury et al., 2017), the effects of genet or substrate on the growth rate of A. palmata in restoration facilities are unknown. To help fill this knowledge gap, we performed an experiment to determine the effect of substrate, genet, and their interaction on the growth and survival of A. palmata microfragments during the first six months of life following fragmentation. We hypothesized that coral will grow faster on commercially available ceramic disks than cheaper in-house made cement disks, and that growth rate would significantly differ among genets.

All corals in this experiment survived and experienced positive growth over the course of the experiment. On average, corals grew from a starting size of 0.57–3.08 cm², an increase in size of 545% (Table 1).

As a whole, colony growth was right skewed with a mean of 0.0125 ± σ 0.0062 cm²/day (0.08774 ± σ 0.0434 cm²/week) and a median of 0.0109 cm²/day (0.0760 cm²/week) (Supplementary Figure 3). There was significant variation among microfragments, with the fastest growing microfragments exhibiting an order of magnitude faster growth than the slowest microfragment (daily growth of 0.0031 vs 0.0348 cm²/day or 0.0217 vs 0.2437 cm^2/week). Growth was also more variable for faster growing genets (Table 1 and Figure 2).

The addition of parent colony (i.e., ramet) as a random intercept significantly improved model fit over a model with no random effects, indicating parent colony had a significant influence on coral growth (Likelyhood Ratio Test, L = 66.46, df = 1, p < 0.0001). This model was also more parsimonious than models which included a ramet effect that varied by genet or substrate (Supplementary Table 1). Similarly, model selection indicated that the optimal model included both genet and substrate as fixed effects. This model outperformed both the genet only model (LRT, L = 6.48, df = 1, p = 0.0055) and the substrate only model (LRT, L = 12.72, df = 3, p = 0.0026). However, adding an interaction between substrate and genet did not significantly improve model fit (LRT, L = 3.25, df = 3, p = 0.1773). Multi-model averaging indicated that out of all possible model configurations, 96.6% of model weights included genet as a factor and 91.2% included substrate as a factor, but only 14.4% of model weights included their interaction (Table 2).

Corals placed on in-house made cement plugs grew at 111.9% the rate of corals placed on mass-produced ceramic plugs (0.0132 vs 0.0118 cm²/day, Table 1 and Figure 2). There were also significant differences in coral growth rates between genets (Table 1 and Figure 2), with corals from genet 14-3 growing at 216% the rate of corals belonging to genet XA.

Numerous factors can impact the outcomes of coral restoration including coral growth, success/survivorship, and cost-efficiency of coral propagation. Many of these factors are controllable by restoration practitioners, and thus represent valuable opportunities to optimize the restoration process. Here, we isolated two such factors, plug substrate and genet, to determine their effects on growth rates of Acropora palmata microfragments. To our surprise, we found that there was a slight, but significant increase in coral growth on cement plugs compared with commercially available ceramic plugs. Additionally, our cost calculations have determined that the ceramic plugs cost more money than purchasing the materials to produce cement plugs in house. Because of this difference, as well as the availability of the cement supplies at various retailers, the use of cement plugs produces an increase not only in coral growth, but also in cost-efficiency and accessibility. This becomes even more significant when considering that nascent coral restoration operations, including those in remote areas or with limited funding or resources, can use widely available materials to produce cost-efficient coral substrates.

Despite the growth and cost advantages of cement plugs, ceramic plugs still hold several distinct advantages. First, ceramic plugs are smoother and more regular in construction. This not only makes it easier to maintain labels on corals, but also makes it easier to hold, clean, and outplant them, as their stems are uniform in length, width, and angle. Ceramic plugs also provide a more uniform growth surface which may reduce variability during experiments. Furthermore, after cement plugs are constructed, they must be soaked for approximately six weeks before they are chemically inert and ready for use, a step that is not necessary when using ceramic plugs. Indeed, corals mounted on cement plugs initially grew slower than those mounted on ceramic, only overtaking ceramic plugs around 90 days into the experiment. While we did not explicitly test the reason for this, it is possible that leaching of cement plugs may have not been complete. This highlights the significant preparation time required when using cement plugs and is of particular importance for seasonal restoration efforts. Nevertheless, our results show that cement plugs made on-site can be a viable alternative to more expensive, ceramic plugs.

Coral growth significantly differed between substrates, but our results also show that substrate effects are dwarfed by the effect of genet. Faster growing genets can be fragmented more frequently, producing greater numbers of potential outplants (Baums et al., 2019). Identifying these fast-growing genets may help speed up the restoration process and success of restoration programs, maximizing output while minimizing cost. Additionally, because colony size is one of the main criteria for sexual maturity in corals, higher growth may allow corals to more quickly reach sexual maturity and increase fecundity at the colony level due to increased numbers of oocyte producing polyps (Álvarez-Noriega et al., 2016). Finally, rapid growth can allow outplants to better avoid size-specific mortality factors, such as corallivory or algae overgrowth (Drury et al., 2017).

Though rapid growth is a desirable coral trait, growth rate alone is not the only important consideration when choosing coral genets. Indeed, genet impacts variables other than growth (Williams et al., 2017; Pausch et al., 2018), and high growth rates may correlate with tradeoffs in other areas such as recovery from thermal stress (Ladd et al., 2017). Since the ultimate goal of most coral restoration is to create resilient, self-reproducing reefs, genets should also be screened for disturbance resilience as well as fecundity when possible. If individual genetic “stress testing” is not possible, coral gardening must be supplemented with restoration activities using coral larvae to increase genetic diversity (Lirman and Schopmeyer, 2016). With high genetic diversity, mass die offs might not be as common as some corals will be more thermally tolerant or disease resistant (Muller et al., 2018).

While the results of this study contribute to knowledge surrounding the efficient growth of A. palmata microfragments ex situ, they are not necessarily indicative of how these microfragments perform as outplants in the wild. Indeed, since genet performance can interact with environmental variables, future work should consider the success and growth rates of microfragments once they have been reintroduced to their natural habitat, and individual practitioners should conduct growth experiments similar to the one described here not only to identify differences in growth rates within their own genet supply, but also to ensure the substrate effect we describe also holds under local environmental parameters. Importantly, if using growth experiments for restoration optimization purposes, it is critical to track growth rates over the same duration that corals are going to be raised ex situ, since growth rates can change over time and may interact with treatment group or season. Finally, there is a need to expand such work to other coral species used for restoration. As reefs continue to face increasing human stressors, coral propagation will continue to be a critical tool in maintaining reef survival and, eventually, restoration. Our study shows that decisions such as which substrate or genets to choose for restoration can have measurable and significant impacts on restoration speed, efficiency, and cost. As coral propagation expands, sharing such best practices will become increasingly more important for coral restoration to become more efficient and effective in the future.

The raw data supporting the conclusions of this article will be made available by the authors, without undue reservation.

SH and RN conceptualized the study. EP and BW performed the experiments. RN analyzed the data. All authors wrote the manuscript.

The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.