On 5 May 2016 after four days larval settlement, strong winds and wave action were impacting the Magsaysay reef sites and the base of the larval mesh net in larval plot 3 had partly detached, so this net was removed to avoid damage to corals and the reef. On 6 May, the larval mesh enclosures were removed from the remaining two larval plots and the three control plots. The 60 settlement tiles from the six experimental plots (ten tiles in each of the three larval enhancement plots and in each of the three control plots) and the additional 12 settlement tiles deployed around the outside of control plots as open reef controls were carefully collected and transferred in seawater to a nearby temporary field station on Tanduyong Island (Figure 1). Tile surfaces were viewed while submerged in small trays under stereomicroscopes illuminated with fibre-optic and LED lights, and numbers of settled coral spat and their locations on tiles were recorded. After monitoring, tiles were returned to their specific locations on the reef plots and reattached onto the tile posts with the correct orientation, to enable repeated monitoring of growth and survival of spat on the tiles.

Initial survival on tiles was monitored after 1 and 2 months post-settlement by collecting the tiles and counting the surviving spat on each tile surface under stereomicroscopes at Tanduyong Island. Tiles were subsequently returned to the reef and reattached to their numbered tile posts. After 10 months, recruits on tiles had grown large enough to census underwater, so survival and growth were monitored in situ at 10, 12, 14, 17, 20, 25, 27, 29, 32, and 34 months post-settlement. The A. tenuis recruits that had settled on the reef surfaces within each of the larval enhancement plots were also visible underwater by 10 months, and their characteristic morphology and subsequent growth into recognisable A. tenuis colonies were used to identify them as originating from larvae that settled on the reef during the larval enhancement experiment. Each recruit and juvenile colony on the reef were mapped and a small, numbered metal tag was permanently attached nearby to facilitate monitoring at the same time as recruits on tiles were monitored.

Colony size was measured using vernier callipers to measure the length (l), width (w) and height (h) of each A. tenuis coral on the recruitment tiles and on the natural reef substrata. The mean planar diameter was calculated from the maximum and minimum diameters measured for each colony. The approximate volume was also measured following a spherical formula EV = πr²h, where r = (l + w)/4 (Shaish et al., 2010). In some cases, larvae settled in close proximity to others and colonies grew together, fusing into a larger chimeric colony. Individual colonies that fused together were still counted and measured separately when polyp demarcation or separation lines were still visible. Where individual corals were indistinct in a fused coral colony, the previous individual count was recorded for survivorship, but measured as a single fused colony for growth.

Evidence of sexual reproduction in all surviving recruits was assessed at 23 and 34 months after settlement, just prior to potential spawning periods after full moons when corals were 2 and 3 years old. Coral reproductive status was examined by carefully breaking up to three small branches to observe whether developing and pigmented oocytes and spermaries were present in broken sections of polyps (after Harrison et al., 1984). Branches were then gently wedged back into the colony to avoid loss of tissues and spawning biomass.

The costs of producing coral recruits including all materials, vessel hire and fuel, diving, labour, larval rearing, and capital costs for larval mesh tents were estimated following Edwards (2010). The average costs per colony were estimated by dividing the total costs by the total numbers of recruits alive at 10 months and at 3 years in the larval plots. Costs were initially calculated in Philippine Peso (PhP) and Australian Dollar (AUD), and converted to United States Dollar (USD) values.

To quantify benthic cover of corals, other benthos and the reef community status prior to larval enhancement, digital images of the experimental plots were taken in April 2016 using a Canon G1-X underwater camera and a 1 × 1 m quadrat frame, and images were analysed using CoralNet (Beijbom et al., 2015). Twenty-five random sampling points were generated for each image and the benthic category underlying each point was identified, resulting in 625 points per plot. Only one adult A. tenuis colony was present inside the reef plots prior to the experiment, growing in control plot 2. Some excess macroalgae within the plots was carefully removed by divers to reduce potential physical and allelopathic effects of algal biomass on larval settlement (Ceccarelli et al., 2018). Monitoring of the reef benthic community status was repeated in March 2019 to quantify changes over time in the experimental plots.

Modified Stationary Point-Counts (Bohnsack and Bannerot, 1986) were used to survey the reef fish assemblages in all plots. Ten-minute surveys were completed in each plot, and all non-cryptic fishes inside the plots were identified to species and recorded (Allen et al., 2005), and the total length of individual fishes was visually estimated (in cm). Monitoring of the reef fish assemblages in all plots was repeated in March 2017, 2018, and 2019 to quantify changes.

The three larval enhancement plots and the three control plots were used as statistical replicates (N = 3). Data from the groups of ten tiles within each plot were averaged to quantify mean initial settlement rates, and subsequent growth and number of surviving recruits at each monitoring period up to age 34 months. Data are reported as mean values ± standard error. Differences in larval settlement patterns on different tile surfaces from the larval-enhanced plots after five days of larval settlement were tested using one-way ANOVA. Tukey’s HSD test was conducted post hoc to determine any significant differences in settlement patterns among tile surfaces. The survivorship of coral recruits on different tile surfaces was analysed using non-parametric pairwise comparison survival tests, based on the Kaplan–Meier function (Lee and Wang, 2003). Significant differences in survival patterns of juvenile corals on natural substrata and on tiles from 10 to 34 months after larval settlement were also tested using the same analysis. Growth rates of juvenile corals on recruitment tiles versus growth rates on natural substrata were compared using one-way ANOVA, and significant increases in growth of juvenile corals through time were determined using repeated measures MANOVA. To determine if the assumptions of ANOVA were met, Shapiro-Wilk normality tests and Levene’s test of homoscedasticity were used on each independent variable.

Benthic community patterns within and among experimental plots were compared both before larval enhancement and three years after the larval restoration, using analysis of similarities (ANOSIM) and PERMANOVA (using PRIMER v6) to test for similarities and significant differences in the per cent benthic cover composition of major benthic categories including live corals, soft corals, sponges, other invertebrates, macroalgae, dead coral covered with turf algae, sand, dead coral and rubble.

Differences in fish species richness and fish abundance between larval enhancement and control plots were analysed using Mann-Whitney U-Test in Statistica ^® software. Fish species composition between larval enhancement and control plots both before and three years after the larval restoration was graphically presented in two-dimensional ordination plots by non-metric multidimensional scaling (nMDS) using the Bray-Curtis measure of similarity (PRIMER v6). Data were transformed to fourth root so that each species contributed evenly to each analysis. Two-way ANOSIM (analysis of similarity) with pair-wise comparisons was conducted to formally test the significant differences between controls and each treatment. Similarity percentage procedure (SIMPER) was also employed to identify the fish species that contributed to the dissimilarities.

A total of 1,617,000 (± 124,910.20 SE) 8-day old larvae and 3,008,000 (± 244,767.50 SE) 7-day old F2 larvae were available from the cultures and these were divided equally into three groups for deployment in the three larval enhancement plots. High initial larval settlement rates were recorded on tiles in the three larval plots that had each been supplied with an estimated 1.54 million A. tenuis larvae. A total of 6,307 settled spat were recorded on the thirty tiles, with a mean of 210.2 ± 86.36 spat per tile (Figure 4A). Coral spat had well-developed primordial skeletons indicating rapid settlement after release onto the reef plots (Figure 4B). Mean larval settlement per tile was highest in larval plot 1 (382.2 ± 116.18), lower in plot 2 (138.2 ± 37), and lowest in plot 3 (110.3 ± 43). Highest mean settlement rates occurred on the sides and bottom surfaces of tiles with lower rates on the top surfaces, but these were not significantly different (F = 0.53, P = 0.61) (Figure 4C). No A. tenuis or other larvae settled on tiles in control plots covered in mesh enclosures (Figure 4A), and no larvae settled on the 12 tiles located around the outside of the mesh enclosures on the control plots, indicating there was no natural recruitment during the five-day larval settlement experiment.

As expected, repeated monitoring of settled spat on tiles showed a Type III survivorship pattern characterised by high rates of mortality during the first two months after settlement and lower mortality from two to ten months (Figure 5A). Recruits grew large enough (1.9 ± 0.26 cm mean diameter) to be visible on tiles and on reef substrata after ten months, enabling in situ monitoring of survival and growth. On the settlement tiles, a total of 24 recruits survived to ten months, including 17 recruits on top surfaces, with an additional 261 A. tenuis recruits found on the natural reef substrata in the larval enhancement plots. Some of the larvae settled close together and after a few months some of these juveniles fused to form chimeras. After 25 and 27 months, a total of 14 fused colonies were recorded with each fused colony comprising of two to seven settlers.

From 10–34 months after settlement, the number of recruits on tiles declined slowly with 13 recruits alive on top surfaces after 25 months, and eight of those colonies surviving at 34 months (log-rank test, χ² = −255.24, P = 0.00, top > sides = bottom; Figure 5B). The mean number of surviving colonies on tiles in each of the three larval plots after 34 months was 2.7 ± 0.67 per plot. Higher numbers of juvenile colonies resulting from larval settlement directly on reef substrata were found, with a total of 212 colonies alive in the three larval enhancement plots after 34 months. The mean number of surviving colonies on natural reef substrata in each larval enhancement plot at 34 months was 70.7 ± 16.83, equivalent to ∼5.7 colonies surviving per m² from larval settlement onto available reef areas in each larval enhancement plot (Figure 6).

Growth of A. tenuis recruits, juveniles and colonies in the three larval enhancement plots was similar on recruitment tiles and natural reef substrata (Figure 7A), with no significant differences in average growth rates of colonies on tiles and reef surfaces (one-way ANOVA, F = 3.29, P = 0.14). Average growth rates of colonies on tiles from 10–34 months post-settlement were 90.8 ± 36.28 cm³ mo^–1, and on reef substrata were 112.6 ± 17.55 cm³ mo^–1. Repeated-measures MANOVA showed significant increases in coral volumes through time both on tiles and reef substrata in the larval enhancement plots (F = 11.14, P = 0.000), with no significant difference in volumes between substrata type (F = 0.92, P = 0.392). At 25 months after settlement (June 2018), average volumes of the fused colonies (2158.8 ± 333.02 cm³ mo^–1) were 61% higher than those of individual colonies (824.1 ± 67.22 cm³ mo^–1).

At 10 months post-settlement when recruits were visible on tiles and on the reef, the 24 recruits on the settlement tiles had a mean diameter of 1.4 ± 0.0 cm, and the 261 recruits on the reef substrata had a mean diameter of 2.4 ± 0.07 cm. At 34 months, the eight recruits on the settlement tiles had a mean diameter of 15.9 ± 3.12 cm, and the 212 recruits on the reef substrata had a mean diameter of 18.1 ± 0.82 cm (Figure 7B), with a size range from less than 5 cm up to > 25 cm mean diameter.

Reproductive condition of all colonies was assessed at 23 months’ age, just prior to the potential first spawning period. Five colonies were gravid with pigmented eggs and well developed spermaries. These colonies ranged in size from 13.0 to 21.0 cm mean diameter (Figure 8A), and included one of the colonies growing on tiles, and four colonies growing on natural reef substrata in two of the larval plots. These colonies were observed spawning from 1830 to 1900 h on 1 and 2 May 2018 (1–2 nAFM, Table 1). Gametes collected from spawn cones placed over the gravid colonies were combined and transferred to the BML hatchery where high rates of fertilization were confirmed.

At 34 months, mean colony size for colonies on both tiles and reef substrata was 17.0 ± 1.86 cm and colony sizes ranged from less than 6 cm to more than 40 cm mean diameter (Figure 8B). A total of 77 colonies were sexually reproductive at this size and age, with gravid colonies ranging in size from 13.2 to 42.3 cm mean diameter (Figure 8B). These gravid colonies included four colonies growing on tiles and 73 colonies growing on natural reef substrata. Gravid colonies were observed spawning (Figure 8C) from 1830 to 1900 h on 23 and 24 May 2019 (4–5 nAFM, Table 1), and gametes were collected from the spawn cones and transferred into floating larval pool nets on the reef for larval culture.

The production cost for each of the 285 sexually derived A. tenuis coral colonies alive at 10 months was United States $13.73. At 34 months age, the production costs for each of the 220 colonies in the restored breeding population was United States $17.79 (Supplementary Table 1).

In 2016, reef benthic cover and coral community status were very similar in the three larval enhancement plots and the three control plots (Figure 9A) with no significant differences in community structure (ANOSIM, R: 0.04, P = 0.30). All plots were degraded and characterised by low mean cover of living scleractinian corals (18.5% ± 2.04%), with similar mean cover of macroalgae and very low cover of soft corals, sponges, other invertebrates and dead coral covered with turf algae, together comprising 21.4% ± 5.72% mean benthic cover. Mean cover of dead coral substrata and coral rubble surfaces potentially available for coral larval settlement was 50.9% ± 6.86%, which represents about 12.5 m² of the reef area within each of the 25 m² plots (Figure 9A).

The 2016 coral community had low mean cover of Acropora spp. and higher cover of encrusting Montipora spp., with low cover of Pocilloporidae, Poritidae, Merulinidae and other taxa (Figure 9B). There were no significant differences in mean cover of Acropora (R: −0.14, P = 0.10), Montipora (R: −0.14, P = 0.60), Porites (R: −0.14, P = 0.10), Pocilloporidae (R: −0.04, P = 0.80) or Merulinidae (R: −0.14, P = 0.10) between larval enhanced and control plots prior to the larval restoration experiment.

In March 2019, coral cover had increased 35 months after larval restoration (Figure 9C), with the larval enhancement plots having significantly higher mean cover than controls (univariate PERMANOVA: F = 14.16, P = 0.0001), primarily due to the 9.5% ± 1.30% increase in A. tenuis mean cover from the larval restoration. Total mean cover of reef corals increased from 19.6% ± 3.12% in 2016 to 40.5% ± 6.13% in 2019 in the larval enhancement plots, and from 17.3% ± 0.94% in 2016 to 26.2% ± 5.01% in 2019 in the control plots. Mean cover of non-scleractinian benthic categories were comparable to 2016 levels except for a reduction in macroalgae cover and hard substrata, corresponding with the increased cover of reef corals (Figure 9C).

Three years after larval restoration, A. tenuis cover had increased significantly in the larval enhancement plots compared with control plots (univariate PERMANOVA: F = 158.88, P = 0.0001) in which no new colonies of this species were recorded (Figure 9D). Mean cover of encrusting Montipora spp. had increased substantially in both larval enhancement (univariate PERMANOVA: F = 189.16, P = 0.0001) and control plots (univariate PERMANOVA: F = 3.73, P = 0.05) (Figure 9D). Mean cover of other reef corals was similar between surveys in 2016 and 2019 in both larval enhancement and control plots (Figures 9B,D).

Surveys of fish assemblages in 2016 showed similar low mean fish abundance and species richness in all experimental plots before the larval restoration experiment due to the degraded status of the reef. There were no significant differences between control and larval enhancement plots for fish abundance (Figure 10A; F = 2.7; P = 0.54) or fish species richness (Figure 10B; F = 1.86, P = 0.70) prior to the larval enhancement experiment. Pomacentridae and Labridae were the most abundant fish families, and had similar mean abundance in the larval enhanced and control plots in 2016, with low numbers of Chaetodontidae corallivores present (Figure 11).

Mean fish abundance and species richness varied slightly between larval enhanced and control plots and among years. Mean fish abundance increased in the larval enhancement plots in 2018 (Figure 10A) and was significantly higher compared to mean abundance in control plots in 2018 (Mann-Whitney U-Test: Z = −1.9640, P = 0.0463). There were no significant differences in fish abundance between larval enhancement and control plots in 2016, 2017, or 2019. Mean fish species richness was slightly higher in larval restoration plots than in control plots in 2018 and 2019 but these differences were non-significant (Figure 10B). The slight increase in fish abundance and change in reef fish assemblages between monitoring years is partly attributable to the increase in Pomacentridae (Figure 11B), specifically the turf farmer pomacentrids such as Pomacentrus burroughi, P. chrysurus, and Plectroglyphidodon lacrymatus.

Reef fish assemblages were similar in all plots, although assemblages in larval enhancement plots were mostly clustered separately from the control plots in each of the monitoring years (Figure 12). However, reef fish assemblages recorded in 2016 were significantly different (R = 0.466, P = 0.001) from the assemblages recorded in 2017, 2018, and 2019 (Figure 12). Pairwise comparisons showed significant differences between 2016 and each of the subsequent years (P = 0.01), but no significant differences between the years 2017, 2018, and 2019. In addition to the turf farmer fishes, pomacentrids that take refuge within Acropora coral branches such as Dascyllus reticulatus, Chromis viridis, and Amblyglyphidodon curacao contributed to the differences in reef fish assemblages in 2016 and later years.

A conceptual diagram summarising A. tenuis recruitment and growth, and changes in coral cover and reef fish assemblages during the three years following the 2016 larval restoration at Magsaysay reef is provided at Figure 13.

Coral restoration should aim to re-establish breeding coral populations on damaged reefs using methods that are cost-effective and scalable, with restored populations capable of surviving and adapting to altered environmental conditions and stressors (Harrison, 2021). However, before restoration is attempted on any reef site it is important to use a decision framework to evaluate the need for such active interventions, and the likelihood of success using appropriate methods.

Initial baseline surveys are important to establish the status of reef communities, and determine the likelihood of natural recruitment enabling recovery within an appropriate timeframe without intervention. If the reef system is degraded and has very low natural recruitment and therefore unlikely to recover naturally, then active intervention is warranted if environmental conditions are potentially suitable for restoring coral populations in a cost-effective manner. Magsaysay reef, chosen for this study, is badly degraded with an algal phase-shifted reef community now characterised by low mean cover of live corals and few natural recruits, as it was in the 2013 larval restoration pilot study (dela Cruz and Harrison, 2017). That study also showed natural recruitment rates were low, and dominated by brooded pocilloporid spat with minimal Acropora recruits present on > 300 recruitment tiles deployed during a two-year period. Therefore, Magsaysay reef is unlikely to recover naturally without intervention.

The decision framework should also consider whether degraded reef systems are potentially recoverable, and the extent to which stressors and key threats that led to coral decline and reef degradation are still operating, or can be potentially managed or tolerated by new generations of restored corals. Previous key threats in the Bolinao-Anda Reef Complex (BARC) where this study was conducted included destructive fishing that is now controlled, a crown-of-thorns starfish corallivore outbreak in 2007 which is unlikely to re-occur given the low coral cover, and intermittent heat-stress and coral bleaching that had minimal impact on the 2013 restored A. tenuis population (dela Cruz and Harrison, 2017). Another key consideration for restoring degraded reefs is the extent to which the reef system is severely phase-shifted with low functional herbivory. Results from the 2013 pilot study showed that although the Magsaysay reef site is algal-dominated, enhancing larval supply significantly increased A. tenuis recruitment on larval restoration plots and restored a breeding population after three years (dela Cruz and Harrison, 2017). Therefore, Magsaysay reef is potentially recoverable, and restoring corals through increasing supply of sexually produced coral larvae can enhance genetic diversity and evolutionary potential, potentially improving environment-genotype matching and resilience of surviving corals (van Oppen et al., 2017).

Our results confirm that mass larval enhancement can significantly increase the initial settlement of corals even on degraded reefs where natural larval supply and coral recruitment have been compromised by loss of adult breeding corals and high turf and macroalgal cover. In this study, we used ∼3.7 times higher supply densities of coral larvae than in the 2013 larval enhancement pilot study (dela Cruz and Harrison, 2017), which resulted in about eight times higher mean initial larval settlement per tile in 2016 compared with 2013. This is consistent with Cameron and Harrison (2020), who found a strong positive relationship between increasing larval density and total larval settlement. Higher rates of settlement in the present study therefore reflect the higher larval supply densities in plots and may also have been enhanced by the use of 7–8 day old larvae that were more fully developed and potentially primed for rapid settlement when released onto the reef areas, compared with the 4 day old larvae used in 2013. Most of the newly settled spat in the 2016 study had well-developed skeletons visible through the translucent polyp tissues (Figure 4B), indicating that the larvae probably settled rapidly after being released into the mesh enclosures on the reef. In addition, the mesh tent enclosures may have allowed larvae to actively swim and search for suitable settlement sites more effectively than under the flat mesh sheets used in the 2013 study. No A. tenuis recruits settled on tiles in control plots covered in mesh enclosures or on the 12 tiles in open control areas without mesh, which confirms low natural larval supply during the larval settlement experiment, similar to previous pilot studies (dela Cruz and Harrison, 2017, 2020b).

Mortality of settled spat and juvenile colonies on tiles was highest during the first 10 months after settlement and then stabilised (Figure 5), consistent with the Type III survivorship curve reported for other broadcast spawning corals (Babcock, 1985; Wilson and Harrison, 2005; Vermeij and Sandin, 2008; Doropoulos et al., 2016; dela Cruz and Harrison, 2017, 2020b) and marine invertebrates (Keough and Downes, 1982; Roughgarden et al., 1985; Hunt and Scheibling, 1997). The causes of mortality of the microscopic newly settled polyps on tiles are not known, but previous studies have noted overgrowth and competition from other benthic biota including allelopathic effects of algae, predation, damage from herbivore grazing, and reduced water quality and runoff from nearby coastal communities contributing to juvenile mortality (Sammarco and Carleton, 1981; Harrington et al., 2004; Penin et al., 2010; Guest et al., 2014; dela Cruz and Harrison, 2017; Cameron and Harrison, 2020). High mortality during early post-settlement life stages may also be a consequence of newly metamorphosed settled polyps having insufficient energy reserves for survival after expending energy and resources for metamorphosis and the onset of skeletogenesis (Harrison and Wallace, 1990), prior to uptake of mutualistic Symbiodiniaceae.

Recruits were visible on tiles and on reef surfaces by 10 months and repeated monitoring showed high rates of survival up to 35 months, consistent with high survivorship patterns of visible Acropora spp. Recruits after colonies reached size-escape thresholds that were recorded in previous studies (Babcock, 1991; Raymundo and Maypa, 2004; Ritson-Williams et al., 2009; Doropoulos et al., 2012; dela Cruz and Harrison, 2017, 2020b). The survivorship pattern of A. tenuis recruits from larvae that settled directly on reef substrata in 2016 (Figure 5) was intermediate between the pattern recorded for A. tenuis recruits in 2013 (dela Cruz and Harrison, 2017) and for A. loripes recruits (dela Cruz and Harrison, 2020b), and substantially higher than for A. digitifera recruits on ceramic plugs outplanted onto reef areas in Palau after 5 and 11 months (Humanes et al., 2021).

Mean growth rates of A. tenuis recruits and juveniles were similar on tiles and on natural reef substrata, and both were higher than for recruits from the 2013 pilot study (dela Cruz and Harrison, 2017). These growth rates were higher than those recorded for A. tenuis settled in a hatchery and then outplanted onto reef areas after 18 months in Akajima, Japan (Iwao et al., 2010), and for A. millepora colonies grown in the BML hatchery or outplanted as sub-adults onto BARC reef areas in Northern Luzon, Philippines (Baria et al., 2012; Guest et al., 2014). Growth rates recorded during the present study were also higher than for A. loripes colonies growing from larvae settled directly in larval enhancement plots on Magsaysay reef (dela Cruz and Harrison, 2020b), and for A. digitifera colonies settled in an ex situ nursery and outplanted at five or 11 months (Humanes et al., 2021).

The rapid growth of most A. tenuis colonies resulting from direct larval settlement on reef areas indicates that environmental conditions are still suitable for survival and growth of this species on Magsaysay reef, despite its badly degraded status. Sea temperatures during this study ranged from 26.5 to 31°C (Supplementary Figure 1) and remained below the coral bleaching thresholds observed previously at Magsaysay reef (dela Cruz and Harrison, 2017). Seasonal changes were also evident in other environmental parameters including periods of reduced salinity and high turbidity from rainfall and runoff associated with monsoon conditions (Supplementary Figure 2). It is possible that elevated organic inputs and nutrients from seepage and runoff from adjacent coastal towns onto these nearshore reef systems are providing supplementary allochthonous food resources that are enhancing heterotrophic particulate feeding and dissolved organic matter uptake (Sorokin, 1993; Anthony, 2000), supplementing energy supplied from photosymbionts and predation on plankton. Larval settlement behaviour and suitable microhabitat selection directly on the reef may also have contributed to rapid growth of surviving colonies in comparison with other Acropora colonies reared from larvae that were settled in nurseries and subject to artificial environmental conditions and selection pressures during early life stages (Iwao et al., 2010; Baria et al., 2012; Guest et al., 2014).

The high densities of larvae supplied to the larval enhancement plots resulted in gregarious settlement of some larvae, and subsequent growth and fusion resulted in 14 chimeric colonies consisting of between two and seven individuals after two years. Chimerism has been reported among populations of Acropora species on reef areas (Puill-Stephan et al., 2009; Schweinsberg et al., 2015) and in experimental studies (dela Cruz and Harrison, 2017; Doropoulos et al., 2017; Cameron and Harrison, 2020; Sampayo et al., 2020), and may lead to enhanced early growth and survival to larger size refugia with potential for increased adaptive potential to changing environmental conditions.

Rapid colony growth also resulted in early onset of sexual reproduction in five colonies (13 to 21 cm mean diameter) that had mature gametes at two years after settlement, and therefore oogenesis was likely to have been initiated about 9 months prior to this at smaller colony sizes (Wallace, 1985; Harrison and Wallace, 1990; Randall et al., 2020). This is the fastest growth to sexual reproduction yet recorded in Acropora corals, and more rapid than predicted based on previous studies of this species (Wallace, 1985; Iwao et al., 2010; dela Cruz and Harrison, 2017) and other Acropora (Baria et al., 2012). Most colonies were sexually reproductive at three years of age with mean colony sizes above 13 cm diameter, thereby re-establishing a functional breeding population on the degraded reef system. The variability in colony sizes among breeding colonies is likely to reflect individual phenotype and holobiont responses to environmental conditions on Magsaysay reef, and highlights the complexity of predicting the age and size of sexual reproduction among reef corals (Wallace, 1985; Harrison and Wallace, 1990; Randall et al., 2020).

The timing of A. tenuis spawning was consistent among years and populations with all colonies recorded spawning during the crepuscular period around sunset, similar to spawning records from the Great Barrier Reef (Harrison et al., 1984; Willis et al., 1985; Babcock et al., 1986) and elsewhere (Baird et al., 2021). The lunar periodicity of spawning varied slightly among years and between the reef and BML culture facility (Table 1). Variability in the lunar night of spawning has been recorded among many Acropora populations in various reef regions, whereas some other taxa such as Merulinidae exhibit more consistent lunar periodicity of spawning, indicating that coral taxa may respond to proximate cues and ultimate selective pressures in different ways (Babcock et al., 1986; Harrison and Wallace, 1990; Hoadley et al., 2016; Randall et al., 2020).

The rapid re-establishment of breeding populations is important for initiating recovery of degraded reefs. Success at localised scales, such as achieved in this study, creates ongoing opportunities to expand larval restoration efforts to adjacent reef areas in future, using some of the millions of gametes now released annually by this population. These breeding populations also contribute to the depleted natural larval supply in the Lingayen Gulf, and some of these genetically diverse larvae are likely to disperse to other reefs and enhance recruitment and reef connectivity at larger scales over time (Harrison, 2006; Jones et al., 2009; Randall et al., 2020). As these breeding colonies grow and their spawning biomass increases, they become more fecund and hence their ecological value increases. In addition, the high densities of breeding colonies established during this reef trial are likely to enhance fertilisation rates on Magsaysay reef from high sperm and egg concentrations following synchronous spawning events (Oliver and Babcock, 1992; Levitan and Petersen, 1995; Yund, 2000). Remarkably, we know very little about the densities of breeding corals required to maximise fertilisation and cross-fertilisation rates on reefs, therefore future restoration trials should consider not only the overall abundance and spatial scales of restored colonies but also their densities for optimising breeding success (Teo and Todd, 2018).

Key issues for coral restoration that need to be resolved are production costs and scalability. The average production costs for each A. tenuis coral colony at 10 months was United States $13.70, and United States $17.80 for each of the 220 colonies in the restored breeding population at 34 months age. These costs were lower than for the 2013 pilot study United States $21.00 per colony at 35 months (dela Cruz and Harrison, 2017) and for A. loripes United States $35.00 at 35 months (dela Cruz and Harrison, 2020b), largely as a result of significantly increased larval supply and higher numbers of recruits and adult colonies surviving in the present study. This indicates that the cost-effectiveness of larval restoration should increase as mass larval production increases for larger-scale delivery onto damaged reefs, as long as long as settlement and recruitment density is optimised to avoid negative density-dependent mortality effects (Doropoulos et al., 2017; Cameron and Harrison, 2020). These production costs per colony for direct larval settlement onto degraded reef areas are substantially lower than for colonies reared in nurseries for extended periods prior to outplanting on reefs. For example, the production costs for 2.5 year old A. millepora colonies initially settled and held in the BML nursery for 7–19 months before outplanting onto reef areas in Northern Luzon, Philippines ranged from United States $284 to $61, respectively (Guest et al., 2014), and for 2.5 year old A. digitifera recruits settled and held in a nursery in Palau for 5 and 11 months before outplanting (United States $227 and $49, respectively; Humanes et al., 2021). These simple production cost metrics do not take into account the growing ecological and socio-economic values of these restored breeding colonies, which provide critical habitats for fish and other reef organisms, and annual increases in fecundity and production of millions of larvae.