Coral reefs globally are ecologically (Wilson et al., 2006), economically (Santavy et al., 2021), and culturally significant (Cinner et al., 2009) to developed and developing countries alike (Laurans et al., 2013). Extants reefs are subject to increasingly frequent and intense disturbances, including climate change-induced factors (Hughes et al., 2017). Major proximate threats to the world’s largest coral reef system–the Great Barrier Reef–include predation by Crown-of-Thorns starfish (Pratchett, 2010), unsustainable fishing practices (Santavy et al., 2021), disease outbreaks (Roff et al., 2011), and bleaching due to climate-induced high temperature stress (Hoegh-Guldberg, 1999). As these disturbances occur more often and with greater intensity, heterogenous coral recovery drives changes in reef composition (Madin et al., 2018). This often reduces the quality of immediate recovery and is correlated with negative long-term trends in biodiversity (Wakeford et al., 2008). Future bleaching events in coming decades will likely surpass the thermal tolerance of essential corals, impeding their continued growth, development, and ecological function (Hoegh-Guldberg, 1999). Given that reefs continue to face these intense conditions, it is critical to understand how reef assemblage will be impacted in the long term.

From 2014-2017, the northern Great Barrier Reef faced a harsh sequence of disturbances–Cyclone Ita in 2014, Cyclone Nathan in 2015, and two massive bleaching events in 2016 and 2017 (Wakeford et al., 2008). The physical damage caused by those cyclones coupled with the large-scale bleaching to devastate reefs around the island. Before these stressors, most Lizard Island reefs were coral-rich (typically >40% coral cover; Done et al., 2010; Johns et al., 2014), but by late 2017 many were close to 4% of live coral cover (Madin et al., 2018).These collective phase shifts away from more structurally secure, complex reefs have had dramatic impacts on fish abundance and species richness (Jones et al., 2004; Wilson et al., 2006). Since changes in composition of coral communities tend to be discrete following these episodic stressors, recovery can occur if a reef has sufficient time between perturbations (Gilmour et al., 2013). However, as the time between these events continues to decrease, phase-shifts become more common, rendering such reefs less able to effectively recover to a pre-disturbance condition (T. P. Hughes and Connell, 1999). The rising prevalence of phase-shifts, together with increasingly frequent disturbances, hinders and, in some cases, completely inhibits coral recruitment and community succession, leaving these potential habitats in a volatile state (Baird and Hughes, 2000).

There are extant examples of other reefs undergoing similar conditions. Following a high urchin mortality event in 1983, grazing was greatly reduced across Caribbean reefs where herbivorous fish populations were overfished, leading to the emergence of alternate stable states in which macroalgae made up the majority of cover in previously coral-dominated areas (Mumby et al., 2007). This weakened the resilience of these reefs against future disturbances that target corals more specifically. Between 1980 and 2001, 286 Caribbean reefs experienced ~17% coral cover reduction in years following hurricanes, with no evidence of recovery to a pre-storm condition for at least eight years post-disturbance (Gardner et al., 2005). This suggests that not only does reef condition decline with more frequent and damaging impacts, but that reef recovery is also being impaired. Caribbean reefs have continued to experience a decline in structural and biological complexity to this day, with Caribbean-wide homogenization of current structures and species (Alvarez-Filip et al., 2009). Similar trends of more homogenized, less healthy reefs have been documented across the same time scale in the Philippines (Licuanan et al., 2019), Indonesia (Wolfe et al., 2020), Papua New Guinea (Berzunza-Sanchez et al., 2013). Each of these reefs have also suffered from overexploitation of resources and poor management decisions, but the negative impacts of cyclones and bleaching cannot be mitigated at the source in the same way. This makes it imperative that we better understand how reefs cope with these large-scale disturbances.

In the immediate wake of these disturbances, researchers characterized the post-impact composition of 13 reef sites around the island in 2017. They found that coral cover across 10 of the 19 sites was below 4%, with only two sites having cover above 40%, the highest cover for most sites prior to the four years’ disturbances (Madin et al., 2018). Researchers predicted reef structure and complexity would begin to return. I sought to document any potential coral recovery six years after the 2017 surveys to determine if that recovery had begun on Lizard Island reefs. In the fall of 2023, I sought to test the hypothesis that coral cover and composition were unchanged from 2017.

I explored coral reefs around the Lizard Island Research Station (14.6688° S, 145.4594° E), on the northern portion of Australia’s Great Barrier Reef. I sampled eight reef sites at Lizard Island from 23 October to 18 November,2023, using the same site name and locations as the previous study (Madin et al., 2018)–Cook’s Path (14.660557° S, 145.453396° E), Lagoon 1 (14.686888° S, 145.456658° E), Lagoon 2 (14.698240° S, 145.450682° E), Horseshoe (14.687851° S, 145.444216° E), Vicki’s (14.685130° S, 145.442876° E), Osprey (14.667579° S, 145.443119° E), Mermaid Cove (14.647248° S, 145.453939° E), and North Reef (14.645044° S, 145.453987° E). These sites are dispersed across the relatively protected side of Lizard Island, leeward of the prevailing Southeast Trade Winds ( Figure 1 ). Risk management concerns (restrictions to remain in range of radio contact and avoid the rougher sea conditions of Lizard Island’s east side) prevented me from collecting data on the Northeast to Southeast of the island and as such I could only sample a subset of the representative sites across the island reefs.

At each site, six 10-meter line-intercept transects were randomly laid out over the reef face at one to two meters deep, with each transect no further than 10 meters apart from the last (Madin et al., 2018). The substrate type beneath the transect tape was recorded to the nearest milimeter, and hard corals were classified by family to determine the percent cover. The categories recorded at each site included soft coral, Acropora, Porites, Pocillopora, encrusting coralline algae, hard dead coral, rubble, sand, and algae. In data visualization, these types were put into the following groups: hard coral (Acropora, Porites, and Pocillopora), soft coral, algae, encrusting coralline algae, and abiotic substrate (hard dead substrate, rubble, and sand). Transect data were analyzed with JMP Pro (v 17). I compared the proportion of total live coral cover by reef, family proportion of total cover by year, grand means of coral cover proportion by year, and grand means of family richness by year using analysis of variance.

In the wake of the 2014-2017 sequence of disturbances, total live coral cover declined dramatically from pre-disturbance conditions in 2014. Some reefs had single percent coverage of live coral by 2017. In the six years since, total live coral cover on all reefs increased dramatically, with six out of eight reefs harboring more live coral cover than in 2014. The remaining two sites (Lagoon 1 and Lagoon 2) show lower coral cover than pre-disturbance 2014 levels ( Figure 2 ). Grand means of live coral cover across my eight sites went from 39.6% in 2014 to 12% in 2017 to 51% in 2023.

The proportion of soft coral and Acropora cover declined heavily, while the proportion of Pocillopora and Porites cover fluctuated more moderately ( Figure 3 ). By 2023 (six years post-disturbance), all groups recovered by various margins, with the proportion of cover of all groups increasing–Acropora by 20.1% (p = 0.049, F = 2.03, df = 7), soft coral by 9.5% (p = 0.025, F = 1.97, df = 7), Pocillopora by 6.8% (p = 0.015, F = 2.49, df = 7), and Porites by 4.2% (p = 0.009, F = 1.42, df = 7). All groups except soft coral recovered to 2014 (pre-disturbance) proportion of cover ( Tables 1 – 3 ).

The grand mean of live coral cover significantly declined from 2014-2017 (mid-disturbance) but rose significantly higher by 2023 (six years post-disturbance) to rise above the pre-disturbance grand mean of live coral cover ( Figure 4 ). The grand mean of family richness declined from 2014-2017 (mid-disturbance) and recovered significantly by 2023, but not to the 2011 or 2014 (immediate pre-disturbance) grand means of family richness ( Figure 5 ). Family richness across all sites improved by 2023 (4.0 ± 0.11; grand mean ± 1 se); however, Lizard Island reefs have yet to reach their pre-disturbance family diversity (4.77 ± 0.58 in 2014).

Acropora dominance relative to total live coral cover may turn out to be a useful metric for forecasting future near-term condition of impacted reefs on Lizard Island. To the extent reef managers can influence reef communities, they should consider emphasizing efforts which promote restoration or recruitment of Acroporids. Recovering these key reef builders seems necessary to recover historic conditions as fast as possible. Before the four-year period of consecutive disturbance events, most Lizard Island reefs were relatively live-coral rich, typically with >40% coral cover (Done et al., 2010; Johns et al., 2014), but by late 2017 many were close to 4% (Madin et al., 2018). Between the intense patches of physical damage from the cyclones and the more uniform impact of the bleaching events, these reefs were devastated, with none of the eight study sites reaching more than 20% total coral cover (Madin et al., 2018). After six years of relatively calmer conditions with fewer conspicuous disturbances, living coral cover at the study sites improved by various margins between 18.4 ± 0.6 to 59.9 ± 5.3%, with many sites on the higher end of that spectrum. Aggregate coral cover across reefs at Lizard Island has recovered since the wake of disturbances ending in 2017, with their rate of recovery closely related to the physical environment they inhabit. More regular (annual or biennial) monitoring of each of these sites could also better inform scientific understanding of the long-term recovery of Lizard Island reefs in the face of continual disturbances at the hands of cyclones, bleaching, Crown-of-Thorns Starfish, and other environmental pressures.