Introduction

Ecological restoration has the potential to reverse ecosystem degradation and restore lost biodiversity and ecosystem services when combined with other natural resource management measures. Indeed, restoration is now seen as increasingly necessary to conserve biodiversity in the face of rapid global change (Possingham et al. 2015; Perring et al 2018). However, despite decades of research and practice, restoration outcomes vary widely, with many projects failing to deliver measurable benefits (Eger et al. 2022; Suding 2011). Setting clear restoration goals and choosing the most appropriate methods to achieve them are both critical to success (McDonald et al. 2016). In particular, the method for deploying new individuals to degraded habitats and the choice of species and life histories have implications for long-term restoration outcomes (Thomas et al. 2014; Bayraktarov et al. 2016).

Ecosystem management paradigms for coral reefs in the Anthropocene are changing due to the impacts of climate change. Proactively tackling the causes of decline by reducing environmental pressures (i.e., reducing carbon emissions) remains the greatest priority (Morrison et al. 2020). However, “reactive” measures, such as reef restoration may also be required to promote recovery under certain circumstances, for example, where poor larval supply limits recruitment of new corals (Anthony et al. 2020; Hein et al. 2021a). The practice of coral reef restoration broadly describes any active conservation intervention designed to assist or accelerate the recovery of coral populations and associated reef structure, function and ecosystem services (Hein et al. 2021a). Despite decades of research and practice, there are only a handful of case studies showing promising outcomes (Williams et al. 2019; Hein et al. 2021b; Peterson et al 2023). This is partly due to a lack of systematic long-term monitoring of restoration projects, but there has also been a focus on methods that can be applied at sub-hectare scales (e.g., reattaching corals of opportunity) and on certain coral taxa and life histories (primarily fast-growing branching species) (Boström-Einarsson et al. 2020; Ferse et al. 2021). Establishing coral reef restoration as a viable management tool in the Anthropocene will require innovations to effectively upscale restoration for a wider range of coral taxa and life histories.

Corals exhibit a spectrum of morphologies and life-history strategies, but many taxa can be categorised into one of four broad life-history groups (stress tolerant, competitive, generalist or weedy) (Darling et al. 2012). Competitive taxa tend to be fast-growing, have branching or plating morphologies, are susceptible to disturbances such as heat stress and predation, but tend to recruit and colonise reef substratum rapidly following disturbances. Stress tolerant taxa, in contrast, grow more slowly, have massive or sub-massive morphologies, tend to be longer-lived and more resistant to disturbances but are slower to recruit compared to competitive taxa (Marshall and Baird 2000; Glassom et al. 2004). The majority of coral reef restoration interventions involve transplantation of competitive coral taxa, and while this approach may yield rapid results, any gains may be short lived. Indeed, in forest restoration, when compared to fast-growing tree species, slow-growing trees accumulate more biomass and contribute more to carbon stock in the long term (Shimamoto et al. 2014). Using slower growing, more stress tolerant taxa in reef restoration may provide a better long-term return on investments, but there are very few reef restoration studies focused specifically on these taxa, making it difficult to assess the potential benefits (Edwards and Clark 1999).

Asexual propagation is the most widely used method for producing coral outplants in reef restoration efforts. These may be from small fragments reared in nurseries or from “corals of opportunity” (Boström-Einarsson et al. 2020; Ferse et al. 2021). In the case of slow-growing massive corals, micro-fragmentation may be used to accelerate early growth rates (e.g. Forsman et al. 2015). While these approaches are simple to implement, they may result in restored populations with low genotypic diversity and they rely on sourcing material from existing healthy colonies. In contrast, sexual coral propagation is increasingly seen as a more scalable approach, because it produces far greater genotypic diversity and enormous numbers of propagules can be harvested from colonies or spawn slicks (Doropoulos et al. 2019; Randall et al. 2020; Vardi et al. 2021). Coral larval propagation may involve seeding competent larvae directly to reef substratum (Heyward et al. 2002; Edwards et al. 2015; dela Cruz and Harrison 2017) or settling larvae onto appropriate substrates for rearing and subsequent outplant (Nakamura et al. 2011; Villanueva et al. 2012; Guest et al. 2014; Chamberland et al. 2017). The feasibility of producing a first filial generation (F1) and rearing these until maturity in situ has been demonstrated for several competitive coral taxa (e.g., Iwao et al. 2010; Baria et al. 2012; Chamberland et al. 2016; Ligson and Cabaitan 2021). While research is ongoing to develop coral larval propagation for other life history groups (e.g., Marhaver et al. 2015; O’Neil et al. 2021), only one study to date has reported long-term outcomes of sexual propagation for a massive coral (Bonilla et al. 2021). None to date have documented the full cycle of rearing and outplanting of slow-growing, stress-tolerant coral species from eggs to spawning adults.

In this study, the massive and sub-massive corals Favites abdita and F. colemani were reared from eggs and larvae, settled on substrates designed to be both easily maintained in nurseries and readily attached to degraded reefs (coral plug-ins). Juveniles on plug-ins were raised in ex situ and in situ nurseries then transplanted to the reef and their growth and survival were monitored until 6-year post-fertilisation. The proportion of colonies that were sexually mature after 5 and 6 years in situ was documented to ascertain the timing of onset of sexual maturity. We estimate the total cost involved in rearing sexually propagated massive corals until maturation for coral reef restoration and project the number of outplants needed to implement restoration at moderate scales.

Materials and methods

Study sites and species

All work was carried out at the Bolinao Marine Laboratory (BML) of the University of the Philippines (see map, Online Resource 1). The merulinid corals F. colemani and F. abdita are hermaphroditic broadcast spawners and have been documented to release their gametes during the major coral spawning period in Bolinao between March and May (Vicentuan et al. 2008; Maboloc et al. 2016). Both species are common on Indo-Pacific reefs and form massive (F. abdita) or sub-massive (F. colemani) colonies and have similar characteristics in terms of their reproductive biology (Maboloc et al. 2016). Details of the study species, coral collection, spawning and larval culture are described in detail in Online Resource 1.

Settlement of coral larvae and early post-settlement survivorship

To facilitate settlement, rearing and subsequent transplantation to the reef, larvae were settled onto coral “plug-ins” (see Guest et al. 2014 for detailed description). Each plug-in comprised a cylindrical cement head (20 mm diameter, 15 mm height, 1492 mm2 surface area) with a polyethylene wall plug (10 mm diameter and 50 mm height) embedded in the base. Plug-ins were biologically conditioned in flow-through seawater tanks for approximately one year, to enable development of CCA and biofilm suitable for induction of settlement (Heyward and Negri 1999). Conditioned plug-ins were cleaned to remove fouling organisms and placed on the bottom of the rearing tanks on the third day post-fertilisation, when settlement competency rates of the two coral species were more than 90% (for methods and results see Online Resource 1). Plug-ins were left in the rearing tanks for 10 days to allow coral larvae to settle and metamorphose into primary polyps. During this period 50% water changes were conducted daily. To estimate early post-settlement survival rates of settled coral spat, all spat were counted on randomly selected plug-ins at approximately 2, 4, 7, 9 and 11-weeks post-fertilisation. On each survey occasion, six plug-ins were selected at random from each tank and from each species (n = 18 plug-ins per species). Different plug-ins were randomly selected on each occasion to avoid repeated measures.

Ex situ and in situ nursery phases

After 10 days, all plug-ins containing at least one settled coral spat (n = 2069) were inserted into six square 1-m2 rearing trays consisting of polyethylene mesh (1 cm diameter holes) supported by PVC pipe frames. Approximately, 350 plug-ins were inserted per frame and transferred to three aerated 400 L flow-through seawater cement tanks for ex situ rearing. Juvenile herbivorous gastropods (Trochus maculatus and T. niloticus) that had been reared three months prior to coral spawning were added to each tank after 2-weeks in order to control algal growth following methods described by Villanueva et al. (2012, 2013).

In May 2010, after 12 months of ex situ rearing, all plug-ins containing a living coral were transferred to an in situ fixed nursery table located at Malilnep channel close to the source site for the colonies (but within a channel protected from strong waves and storms) (Online Resource 1).

Transplantation and monitoring of reared corals

To test the potential for outplanting reared corals on plug-ins, sites were selected based on the following criteria: (a) the two coral species occurred naturally at the sites, (b) coral limestone outcrops suitable for transplantation were present, and (c) the sites were at a similar depth to the donor sites (2 m to 8 m deep). Based on these criteria, three sites within the Bolinao-Anda reef complex at varying distances from the initial donor site were selected Lucero (16°24ʹ45ʺ N, 119°54ʹ15ʺ E), Marcos (16°18ʹ0ʺ N, 120°01ʹ20ʺ E) and Caniogan (16°16ʹ40ʺ N, 120°00ʹ40ʺ E) (Fig. S1).

An initial pilot outplantation was carried out in October 2010 (17 months post-fertilisation) with 30 plug-ins from each species at one site (Lucero) to test the attachment method and early survivorship. The pilot outplant proved successful with 83–87% of plug-ins containing a living coral 8-months post-outplant (Online Resource 1). Following the success of this pilot, a larger outplantation was initiated in June 2011 (25-month post-fertilisation) by selecting the 120 plug-ins containing the largest live corals from each species from the in situ nursery for transplantation to the fringing reefs off Lucero, Caniogan and Marcos. At each site, four outcrops (1.5 to 3 m diameter and 2 to 6 m depth) were selected and their locations were mapped. The reef substrate was brushed to remove sediment, algae and other fouling organisms. Twenty holes at least 10 cm apart were drilled on each of the four outcrops using a pneumatic drill attached to a scuba tank, and 10 plug-ins with F. colemani and 10 with F. abdita were deployed on each outcrop (one coral plug-in per hole), resulting in 40 plug-ins of each species at each of the three sites (n = 240 plug-ins in total). The corals were tagged by attaching a stainless-steel label to the base of the coral and the positions of the colonies on each plug-in were mapped. The holes were drilled deep enough to ensure that the cement heads of the coral plug-ins were in contact with the substrate and plug-ins were stabilised with marine epoxy.

Assessment of plug-in yield, coral radial growth and reproductive maturity

Assessments of survivorship and growth were carried out in different ways during the ex situ nursery phase (0–12-month post-fertilisation), the in situ nursery phase (13–25 months) and after outplantation (26–71 months). The number of coral plug-ins with a live coral (i.e. the yield) was used as a measure of survivorship. This is because, from a restoration perspective, the unit of interest is the number of outplanted substrates that yield a live coral, rather than the proportion of surviving spat (Guest et al. 2014). Therefore, after 6 months post-fertilisation, only the total number of plug-ins containing at least one living coral was counted and not the total number of colonies. Plug-in yield was monitored at approximately 6- and 12-months post-fertilisation during the ex situ nursery phase, at 17- and 25-months post-fertilisation during the in situ nursery phase, and at 30, 40 and 72 months post-fertilisation following outplant to the reef.

Growth measurements on individual corals began at 12 months, when plug-ins were transferred to the in situ nursery. Coral radial growth rates in the nursery were estimated by measuring the geometric mean diameter (GMD) as follows:

$$\mathrm{GMD}= \sqrt{D1 \times D2}$$

where D1 is the maximum diameter and D2 is the perpendicular diameter. GMD was estimated from corals on 50 plug-ins of each species (each containing only one coral colony) in the nursery. Annual radial growth rates were calculated as the difference between the final and initial geometric mean radius (GMR), divided by the number of years of measurement. For the nursery corals, GMD was measured on six occasions at approximately 12, 15, 17, 23 and 25-months post fertilisation. During the outplant phase, coral GMD of all surviving outplanted corals was measured at approximately 29, 40 and 72-months post-fertilisation. During the outplant phase, corals had completely overgrown the plug-ins, so in addition to diameters D1 and D2, height (H) was also measured to obtain 3-dimensional colony radius (3D GMR) as follows:

$$\mathrm{GMR}= \sqrt[3]{\frac{D1}{2}\times \frac{D2}{2}\times H}$$

and these were used to estimate radial growth in mm yr−1.

In May 2014 (60 months post-fertilisation), surveys were conducted at two of the outplant sites (Lucero and Caniogan) to estimate the proportion of colonies that were gravid. The transplants in Marcos were not sampled due to logistical constraints (poor weather conditions). The outcrops were haphazardly selected and between three and six colonies of each coral species on each outcrop were sampled (n = 41 sampled in total). Several polyps were removed using a hammer and chisel, and the presence or absence of pigmented oocytes was noted in situ as this usually indicates spawning on or after the next full moon. A second, more comprehensive sampling was carried out in April 2015 (~ 72 months post-fertilisation) at all three sites (data for F. abdita previously reported in Bonilla et al. 2021). All surviving colonies from the transplantation carried out in 2011 (n = 188) were sampled to record the presence or absence of pigmented oocytes for both species.

Data analysis

To model the early post-settlement mortality, a standard negative exponential survival curve was fitted to the first three points (2–7 weeks post-fertilisation – shown as filled circles) which covered the period of decline:

$${N}_{t}= {N}_{0} \times {e}^{-Mt}$$

where Nt is the number of recruits surviving at time t, N0 is the notional number of initial settlers, e is the exponential and M is the instantaneous coefficient of mortality. Counts of live coral spat on plug-ins were non-normal and heteroscedastic. To test for differences in early post-settlement mortality for the final three time points for both species, ANOVA followed by Tukey’s test was carried out on counts of coral spat. Normality and homogeneity of variances were achieved for each species by a logarithmic transformation of the counts prior to analysis. For corals outplanted to the reef, a two-way ANOVA was carried out to test for differences in growth rates (3D GMR mm yr−1) between species and among sites, with species and sites as fixed factors in the analysis.

Cost estimates and scaling exercise

The methods for estimating costs are described in detail by Edwards et al. (2010) and Guest et al (2014). The cost of equipment and consumables were separated from labour costs, boat hire and scuba tank hire. Wage rates were based on standard local wage rates at the time the work was carried out in the Philippines between 2009 and 2015. Different wage rates were set depending on the skill level required for each task, based on wage rates during the time the study was carried out, as follows: level 1 (highest salary, e.g. scientific adviser/expert), US$5.63 h−1; level 2 (medium salary, e.g. trained local staff), US$3.50 h−1; and level 3 (lowest, e.g. trained manual labour), US$1.31 h−1. We use local wage rates as it is likely that any restoration effort would need to be carried out by locally based practitioners for it to be sustainable. Labour is estimated as person hours so costs could be converted to wage rates in different countries. Costs for capital equipment, i.e. equipment such as microscopes and diving equipment that could be used for longer than the duration of this project, were given a three-year life span, therefore the total capital equipment costs were divided by three. The cost per coral was estimated by dividing the total project cost by the number of plug-ins with at least one live coral for different production stages: (1) a coral in the ex situ nursery after one year, (2) a coral transplanted to the reef after 1 year in the in situ nursery (ca. 2 year old transplant) and (3) the cost for a surviving six-year-old coral transplant (Online Resource 2). Costs per coral can be reduced by producing at larger scales due to economies of scale, therefore a scaling exercise was also carried out to estimate costs if the initial number of plug-ins settled was increased by a factor of five to 10,000 (Online Resource 2). All costs were estimated in US dollars using a conversion rate of 45 Philippine Pesos to US$1 based on the average exchange rate during the study period (2009–2015).

Restoration outcome projection

To estimate the outcome of outplanting corals as part of a real restoration effort, we used growth rates and mortality rates from the present study to estimate the number of outplanted corals required to a) increase coral colony density of adult colonies by 1 colony per m−2 over a six-year period in 1 hectare of reef, and b) increase coral cover by 1% over a ten-year period in 1 hectare of reef. One hectare was used as most restoration efforts range in scale from sub-hectare up to a few hectares in scale (Online Resource 3).

Results

Early post-settlement survivorship.

The initial mean coral spat density on plug-ins (Fig. 1) was high for both F. colemani (538 ± 70 spat plug-in−1; mean ± S.E.) and F. abdita (468 ± 68 spat plug-in−1) at two-weeks post-fertilisation (n = 18 plug-ins per species). This equates to average spat densities of ~ 36 and 31 spat per cm2, respectively. Although analysis of variance on the transformed counts showed significant differences (P < 0.001) among surveys for both species, Tukey pairwise comparisons showed no significant differences among counts between 7 and 11 weeks. However, for both species there was a significant decline in recruit numbers between 2 and 4 weeks post-fertilisation. There was a significant fit to a standard negative exponential survival curve for both species (F1,52 = 30.17 and 54.79 for F. abdita and F. colemani, respectively, P < 0.0001) over the first 7 weeks after settlement. The half-life (time to 50% mortality) of F. abdita spat was estimated at 2.2 weeks (M = 0.31 wk−1), whereas that for F. colemani, recruits was 1.4 weeks (M = 0.51 wk−1) during the period of high attrition in the first 7 weeks after settlement. At the end of this high attrition period, spat densities were approx. 6.1 and 3.8 per cm2 for F. abdita and F. colemani respectively.

Fig. 1
figure 1

Early post-settlement survivorship of coral recruits showing the mean number of live (a) Favites abdita and (b) F. colemani coral spat per cm2 (± 95%CL of the mean). A regression has been fitted to the first three points to show apparent exponential mortality during the first 7 weeks post-fertilisation. ANOVA of log-transformed data shows no significant change in numbers between 7 and 11 weeks. The greyed area to the right indicates period when there was no significant decline in spat density. Red line is mean spat density at the end of high attrition period

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Coral growth rates in the nurseries and after transplantation

After 12 months in the ex situ nursery, F. abdita and F. colemani on coral plug-ins had average GMDs of 12.1 ± 5.1 mm (mean ± SD) and 14.9 ± 3.9 mm, respectively (n = 50) (Fig. S4). Assuming a mean starting diameter of 0.4 mm (Maboloc et al 2016), radial growth rates during this period were equivalent to 5.8 mm yr−1 and 7.3 mm yr−1, respectively, for F. abdita and F. colemani. During the subsequent 13.3 months of measurement at the in situ nursery, corals increased in diameter by 2.1 and 1.4 fold and had average GMDs of 25.9 ± 3.5 mm and 21.4 ± 5.9 mm (mean ± SD), respectively, for F. abdita and F. colemani (Fig. S4). Geometric mean radial (GMR) extension rates during this period were equivalent to 6.2 mm yr−1 and 3.0 mm yr−1, respectively, for F. abdita and F. colemani with the latter’s growth apparently more constrained by the width of the plug-in (Fig. S4). At 71 months post-fertilisation, outplanted corals had mean GMDs of 66.3 ± 22.4 mm and 51.8 ± 15.8 mm (mean ± SD)(Fig. S4). For corals transplanted to the reef, 3D GMR growth rates were 5.0 ± 0.3 mm year−1 (mean ± SE) and 2.9 ± 0.4 mm year−1 (mean ± SE) for Favites abdita and F. colemani respectively. GMR growth rates differed significantly between species (F = 20.3, P < 0.0001) but not sites (F = 0.60, p = 0.549).

Survivorship of coral plug-ins in the nurseries and after transplantation

All of the plug-ins (n = 2069) had at least one live coral spat 10 days after they were introduced to competent larvae. Of these, 1059 plug-ins had F. colemani spat and 1010 plug-ins had F. abdita spat. In the subsequent 12 months at the ex situ nursery the number of plug-ins containing a living coral (i.e., plug-in yield) declined by 34% and 39% for F. abdita and F. colemani, respectively (Figs. 2, 3). During the subsequent 13 months at the in situ nursery, plug-in yield declined by a further 11% and 7% respectively. This meant that after the first 25-months of nursery rearing, there were 535 and 552 plug-ins with living corals remaining, or a plug-in yield of 55% and 54% for F. abdita and F. colemani, respectively (Figs. 2, 3). Survivorship was high for both species following outplantation of a subset of plug-ins (N = 120 per species) from the in situ nursery to the reef. During the almost 4-year period (46.3 months) post-outplantation, plug-in yield declined by 10 and 14% for F. abdita and F. colemani, respectively (Figs. 2, 3). This resulted in a final estimated plug-in yield of 45 and 40% for F. abdita and F. colemani 6-years post-fertilisation. Post-outplantation yield was highest at Caniogan (90 and 88% for F. abdita and F. colemani respectively) and lowest at Marcos and Lucero at 77% at both sites for F. abdita and 70 and 67%, respectively, for F. colemani. Detachment rates were very low with only one plug-in being recorded as detached at 15 months post-outplantation.

Fig. 2
figure 2

Plug-in yield (%) for F. abdita (black line) and F. colemani (grey line). Grey shading indicates the different nursery periods. Numbers on each line are % yield at times surveys were carried out

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Fig. 3
figure 3

Representative photographs of sexually propagated Favites abdita (a-d) and Favites colemani (e–h) in the ex situ nursery 6 months post-fertilisation showing co-reared Trochus niloticus (a, e), at the in situ nursery at 18-months post-fertilisation (b, f), and colonies outplanted to the reef at 40 (c, g) and 71 months post-fertilsation (d, h). Scale = 1 cm

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Assessment of reproductive status

A total of 18 F. abdita colonies (9 at each site) and 23 F. colemani colonies (11 at Lucero and 12 at Caniogan) were sampled to assess reproductive status in 2014. Geometric mean diameters of the F. abdita colonies sampled ranged from 37 to 86 mm but none of the colonies sampled were gravid. In contrast, 90% of F. colemani colonies sampled were reproductively mature. The gravid colonies (n = 21) ranged from 25 to 65 mm in geometric mean diameter, while the two non-gravid colonies (n = 2) had GMDs of 34 and 39 mm. More extensive sampling was carried out in 2015 and all surviving corals at the three sites were sampled to estimate the proportion that were reproductively mature (n = 188). The majority of colonies for both species contained visible pigmented eggs (Table 1, Fig. 4). Non-gravid colonies had GMDs ranging from 16 to 76 mm for F. abdita and between 14 to 43 mm for F. colemani. Data for F. abdita from 2015 were previously reported in Bonilla et al. (2021).

Table 1 Sampling of colonies in 2015 to estimate the proportion of mature and empty colonies
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Fig. 4
figure 4

Six-year-old sexually propagated Favites abdita (a) and Favites colemani (b) transplants showing fragmented polyps and pigmented mature gametes

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Cost estimates

We estimated costs based on the survivorship observed in this study, with a starting production scale of 1000 plug-ins per species. The total expenditure incurred for rearing corals from eggs to six-year-old outplants on the reef was estimated at US$6262 per species (Online Resource 2). Cost per coral plug-in ranged from just over US$4 for a 1-year old juvenile in the ex situ nursery to around US$14 for a 6 year old, sexually mature colony (Table 2). Scaling up production to 10,000 plug-ins reduced costs of six-year-old outplanted corals by ~ 50% (Table 2) (Online Resource 2).

Table 2 Estimates of cost per coral plug-in with at least one live coral at different stages and production scales
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Estimated effort required to achieve restoration goals

We estimated the number of colonies required to a) increase coral colony density of adult colonies by 1 colony per m2 over a six-year period in 1 hectare of reef, and b) increase total coral cover by 1% over a ten-year period in 1 hectare of reef (Table 3). To achieve objective a) based on yields post-outplant reported here (84 and 74%), a total of 11,848 and 13,514 2-year-old corals of F. abdita and F. colemani, respectively, would need to be outplanted to achieve an increase in colony density of one adult per m−2 after 6 years. To achieve objective b), based on the growth rates reported here, it would require 13,284 surviving outplanted colonies of F. abdita to increase coral cover within a hectare of reef by 1% after ten years, whereas 25,423 colonies of F. colemani would be required. Based on survivorship found in this study, you would need to outplant 19,309 coral plug-ins with F. abdita and 52,965 coral plug-ins with F. colemani to achieve this goal (Table 3) (Online Resource 3).

Table 3 Data used to estimate the number of coral outplants required to (a) increase coral colony density of adult colonies by 1 colony per m2 over a six-year period, and (b) increase coral cover by 1% over a ten-year period
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Discussion

Despite the potential advantages of using slow-growing, stress tolerant coral taxa as part of reef restoration efforts, few studies have examined their long-term performance following transplantation. Only one study to date has reported onset of reproductive maturity in a sexually propagated massive coral (Bonilla et al 2021), and none to date have documented the full cycle (egg to spawning adult) of rearing and outplanting stress-tolerant coral taxa using larval propagation. Our study, therefore, represents an important milestone in the development of effective practice for reef restoration. Due to high post-transplantation survivorship, we attained exceptional post-outplant yields relative to similar studies using fast-growing coral taxa at the same location. For both species, outplants were reproductively mature by six years of age, demonstrating the feasibility of creating viable restored populations within a decade. Costs were comparable to those found for other species using sexual propagation and larval seeding, and could be reduced further by increasing the scale of production. Nonetheless, large numbers of outplanted corals would be needed to achieve meaningful restoration outcomes, in part due to relatively slow growth rates of the study species.

Broadcast spawning corals are at their most vulnerable during the earliest stages of development both pre- and post-settlement. Pre-settlement mortality can be reduced dramatically in ex situ systems by optimising water quality and propagule density (Guest et al. 2010; Pollock et al. 2017), however post-settlement mortality can still be high even under controlled conditions (Craggs et al. 2019). Some of the processes driving early mortality may be density dependent (Edwards et al. 2015; Doropoulos et al. 2017), thus it is important to determine optimal starting settlement densities in order to maximise yields in coral larval propagation (Cameron and Harrison 2020; Humanes et al. 2021). Despite high initial settlement densities, the number of live F. abdita and F. colemani spat decreased by 7.5 and 11 times over the first nine weeks post-settlement. Spat of F. colemani, which were more densely settled at the start, died faster than those of F. abdita, suggesting density dependent mortality. While we did not set out to assess the optimal settlement densities for these species, those observed at 7-weeks post-settlement (3.8 to 6.1 spat per cm2) might provide a guide for optimal start densities for massive corals. These spat densities are higher, but of a similar magnitude to those found to be optimal for Acropora digitifera (1.3 per cm2) (Humanes et al. 2021), suggesting higher initial settlement densities are needed for merulinid corals that produce smaller spat compared to Acropora (e.g., Bonilla et al. 2023). Our data suggest that a minimum of ~ 4 to 6 spat per cm2 provides a reasonable starting point suggesting it would be possible to seed up to 20,000 coral plug-ins from a culture of 1 million coral larvae.

Estimating growth rates of corals during the nursery rearing phase was challenging, because growth was constrained by the shape and size of the coral-plug-in. Corals tended to grow laterally until they had reached the width of the plug-in, but then radial growth rates appeared to slow as tissue spread down the sides of the plug-in. Therefore, we base our colony growth rate estimates only on growth measured post-outplant as by that time, corals had completely sheeted over the cement head of the plug-in. Radial growth rates for both species were relatively slow (average ~ 5 and 3 mm year−1) and at 6-years, outplanted corals had average diameters of ~ 6.6 cm and 5.2 cm for F. abdita and F. colemani respectively. This is considerably slower growth (~ 5—19 times) compared to Acropora millepora colonies outplanted at the study location using the same method (e.g., radial growth estimates from ~ 28 to 57 mm year−1) (Guest et al. 2014). Growth rates did not differ among outplant sites but were significantly different between species, with F. abdita having approx. 70% faster radial growth rates. While F. abdita grew faster, colonies were not sexually mature until six years post-fertilisation, compared to F. colemani which were mature at five years post-fertilisation, suggesting a trade-off between growth and reproduction between species.

Yields at 6-years post-fertilisation were comparatively high with an estimated 40 to 45% of initially settled plug-ins containing a living coral. Most of the reduction in yield occurred during the first 13-months in the ex situ nursery (34 to 39%), with the remaining occurring over the subsequent 5 years in the in situ nursery and at the outplant sites. Yields in the ex situ nursery were considerably higher than those reported for Acropora valida juveniles (~ 10% of settled coral substrates with a living coral after six months) reared in a similar way at the study location (Villanueva et al. 2012). Several studies have highlighted the importance of grazers to overcome early mortality bottlenecks (Omori et al. 2006; Toh et al. 2013; Villanueva et al. 2013; Craggs et al. 2019). Co-rearing with grazing gastropods in our study undoubtedly contributed to high yields in the first 12-months because grazers were able to remove filamentous turf algae that would otherwise have smothered juvenile corals. This reduced the need for manual removal (although non-filamentous alga still needed to be removed by hand), highlighting the need for effective grazing and co-culturing in the early stages of coral larval propagation (Craggs et al 2019). It should be noted that yield estimates may be slightly inflated as the largest corals from the nursery were non-randomly selected for outplant.

Comparing yields of outplanted corals among studies is difficult because of variations in nursery rearing periods and study durations, however, estimated annual mortality rates in our study (4.6 to 6.4% per year) were three to twenty times lower than reported for sexually propagated Acropora species at the same study location (ranging from 22 to 99% per year) (Online Resource 4). In these studies, annual mortality of outplanted Acropora ranged from 22 to 93% per year (Omori et al. 2008; Boch and Morse 2012; Villanueva et al. 2012; Guest et al. 2014; Chamberland et al. 2015; Baria-Rodriguez et al. 2019; Ligson and Cabaitan 2021). At the nursery, plug-ins were manually cleaned every fortnight, however no husbandry or maintenance was carried out on corals outplanted to the reef. Their relatively high survivorship, therefore, is largely due to the lower susceptibility of these taxa to stress (e.g., high heat tolerance, lower palatability to predators, greater resistance to physical disturbances) (Darling et al. 2012). One of the strongest arguments against small scale bioengineering solutions (such as reef restoration) is that restored corals are likely to die unless underlying causes of coral decline are addressed (Morrison et al. 2020). Without doubt, addressing causes of coral mortality should be the first priority, however, small scale restoration interventions may be successful if appropriate stress tolerant taxa or phenotypes are selected (Humanes et al. 2021, 2022). For example, in our study, prolonged periods of heat stress > 4 degrees heating weeks (DHW) occurred from June to November 2010, June to August of 2012 and 2013 (NOAA), including a 3-month period of DHWs > 8 in 2010, but these events did not result in significant mortality of nursery or outplanted corals.

Understanding the size and age at which outplanted corals become reproductively mature is critical because production of spawning colonies is a key measure of restoration success (Edwards and Gomez 2007; Bonilla et al. 2021). Colony age at onset of reproduction for slow-growing massive corals is largely unknown (Bonilla et al. 2021, 2023), however, estimates based on colony sizes and known growth rates for several massive species suggest reproductive onset from four to eight years (Babcock 1986). Our study provides empirical evidence supporting these estimates with onset of reproductive maturity at 5- or 6-years post-fertilisation. This is later than for faster growing Acropora species which can mature as early as 2–3 years post-fertilisation (Baria et al. 2012; Harrison et al. 2021). Our results highlight the need to carefully consider life history traits of chosen taxa for restoration efforts to achieve stated goals. For example, while slow-growing taxa may have longer survival times than competitive taxa, fast-growing taxa will contribute to the larval pool earlier and therefore make a greater initial contribution to recovery and recruitment (Bonilla et al. 2023).

Based on production scales used in this study (~ 2000 initially settled plugs) we estimated that it would cost around US$14 to produce each outplanted surviving 6-year old coral. In comparison, reported costs to produce one outplanted sub-adult Acropora (from 12–30 months old) using similar methods at the same study location range from US$11–61 per colony (Villanueva et al. 2012; Guest et al. 2014; Baria-Rodriguez et al. 2019; Ligson et al. 2020). The costs reported here and in other studies are experimental costs, therefore should not be used to estimate actual restoration costs because experimental trials are at smaller scales than real reef restoration interventions (Humanes et al. 2021). When we estimate costs at more realistic, larger scales of up to 10,000 coral plug-ins, the cost per surviving Favites coral drops to ~ US$7 due to economies of scale. This is similar to the cost estimated for 3-year old outplanted Acropora millepora if 5000 plug-ins are produced annually for a 3-year period (Guest et al. 2014). Conceivably, these costs could be reduced further if volunteers are used or if materials and equipment were provided as in-kind support.

A common misconception is that nursery rearing increases costs of coral larval propagation (Harrison et al. 2021), but our estimated costs to produce 6-year old mature merulinid corals (~ US$14) are within the same order of magnitude as those reported to produce 3-year old mature Acropora corals (~ US$18- US$35) using direct larval enhancement approaches (i.e., where larvae are seeded directly to substratum, without any nursery rearing) (dela Cruz and Harrison 2017, 2020; Harrison et al. 2021). This is despite the additional husbandry required during rearing and the labour involved in outplanting individual corals. This is due to a trade-off between survival and the amount of husbandry investment prior to corals being outplanted or seeded. For example, only ~ 0.005% of directly seeded Acropora larvae survive to adulthood (dela Cruz and Harrison 2017, 2020), whereas survival from larvae to outplanted adult corals is an order of magnitude higher (~ 0.05–0.06%) for corals reared in a nursery based on data from the present study.

Based on the growth rates and survivorship in this study, we estimate that between 11,848 and 13,514 2-year-old corals would need to be outplanted to achieve a modest increase in colony density of one adult per m−2 after 6 years. Similarly, to increase coral cover within a hectare of reef by 1% after ten years, you would need to outplant over 19,000 coral plug-ins with F. abdita and almost 53,000 coral plug-ins with F. colemani. These results show that while using stress tolerant corals for reef restoration provides a long-term return on investment due to high survival rates of outplants, considerable time and labour would be required to achieve even modest restoration goals due to slow growth rates. This highlights the need to diversify the range of coral life-history traits and species used in reef restoration with a mix of both fast- and slow-growing taxa. Furthermore, research into methods to improve survivorship of fast-growing taxa for reef restoration interventions should be accelerated. The 200 km2 coral reefs in the vicinity of Bolinao-Anda (where this work took place) had an estimated total economic value of US$38 million per year at the time this study took place (Cruz-Trinidad et al. 2011). This equates to US$1900 per hectare per year. In contrast, we show here that the cost of even modest restoration gains at hectare scales is likely to cost 10’s to 100’s of thousands of US$ using the costs per coral estimated above. This highlights the need for careful consideration to ensure investment in restoration is justified and carries meaningful socio-ecological returns and requires that sustained financing structures are clearly conceived in advance (Suggett et al 2023).

At the onset of this study, the coral plug-in was relatively novel as there were few other devices designed specifically for settling, rearing and directly outplanting sexually propagated corals (e.g. Omori & Iwao 2009). Since its development, there has been concerted research effort to improve devices specifically for sexual coral propagation (Randall et al. 2020). Considering that the main bottlenecks for sexual coral propagation are a) efficiency of outplant and b) post-outplant survivorship, research and development should focus on reducing the time it takes to outplant each device and design modifications that improve survivorship of outplanted juvenile corals. Improvements could be achieved by developing or modifying attachment devices such as the Coralclip® (Suggett et al. 2020), improving the design of devices that “self-attach” (e.g. Chamberland et al. 2017), and designing devices in a way that reduce mortality from fish predation (e.g. Randall et al. 2020; Crawford et al. 2022).

As with all reactive management interventions, the approach tested here would require large investments to achieve even modest gains. Before embarking on costly restoration efforts, it is important to know if they provide a better return on investment compared to other management strategies. It is also critical to ensure that reactive measures (such as coral transplantation) do not distract from dealing with the root causes of decline (e.g., climate change, poor water quality etc.). Nonetheless, our study shows that rearing and outplanting stress tolerant coral taxa in addition to other life-histories, may provide longer-term restoration returns than using fast-growing taxa on their own.