The loss of top predators has cascading effects and has led to long-term shifts in the structure and function of many ecosystems worldwide (Estes et al., 2011; Ripple et al., 2014). Consequently, the protection and restoration of predator populations is considered an important management action to restore trophic interactions, increase ecosystem stability and resistance to climate change impacts (Ripple and Beschta, 2012; Ripple et al., 2014; Sala and Giakoumi, 2018; Peleg, et al., 2023). However, climate change can lead to the arrival of new species through range extensions (Ling et al., 2009a; Poloczanska et al., 2013; Pratchett et al., 2014; Vergés et al., 2014; Pecl et al., 2017) and irruptions (large increase) of native species (Carey et al., 2012; Greenville et al., 2012; Sambaraju et al., 2019) that can alter ecosystem dynamics and lead to novel trophic pathways (Edwards and Richardson, 2004; Miranda et al., 2019). These irruptions are generally facilitated by interactive stressors that reduce overall ecosystem resilience, making them more prone to stress and concurrent outbreaks (Breitburg and Riedel, 2005; Purcell, 2012; Ban et al., 2014; Ling et al., 2015). While the protection of predators is expected to limit irruptions of both native and introduced species (Ling et al., 2009b; Wallach et al., 2015), predator populations and their recovery following protection can also be influenced by climate change (Eisenlord et al., 2016; Harvell et al., 2019) and the extent of historic population depletion (Hutchings and Reynolds, 2004; Neubauer et al., 2013). Thus, in a changing environment, it is becoming increasingly difficult to predict how food webs and ecosystems will respond to protection (Wallach et al., 2015; Rilov et al., 2020; Hammerschlag et al., 2022).

In temperate marine environments, the overharvesting of predators on rocky reefs has long been attributed to the loss of kelp forests in many regions worldwide (Steneck et al., 2002; Filbee-Dexter and Scheibling, 2014; Ling et al., 2015). The loss of predators often triggers a trophic release of herbivorous sea urchins, leading to the deforestation of kelp forests, leaving large areas of reef devoid of kelp (termed “urchin barrens”) with reduced productivity, biodiversity and function (Jackson et al., 2001; Steneck et al., 2002; Filbee-Dexter and Scheibling, 2014). These urchin barrens represent alternative stable states, maintained by strong feedback mechanisms that can persist for decades and be extremely difficult to reverse (Filbee-Dexter and Scheibling, 2014; Ling et al., 2015). Nevertheless, in many cases the protection of predators and a concomitant population increase has resulted in a reversal of this pattern with the recovery of kelp forests (Estes and Palmisano, 1974; Babcock et al., 1999; Shears and Babcock, 2002; Shears and Babcock, 2003; Sangil et al., 2012). This was observed in Alaska where the protection and subsequent recolonization of sea otters (Enhydra lutris) resulted in significant declines of sea urchins and a reforestation of kelp (Estes and Duggins, 1995). Similarly, within no-take marine protected areas (MPAs), the recovery of large predatory fishes and rock lobster has resulted in the reverse shift from urchin barrens to kelp forest following decades of protection (Babcock et al., 1999; Behrens and Lafferty, 2004; Edgar et al., 2009; Eger and Baum, 2020; Peleg et al., 2023). Consequently, the protection of predators in no-take MPAs is widely considered one of the most effective management tools for restoring trophic interactions and reversing the impacts of fishing on kelp forest ecosystems (Sala and Giakoumi, 2018).

The protection of predators within MPAs is expected to increase the resilience of kelp forests and mitigate the impacts of climate change (Ling and Johnson, 2012; Gallardo et al., 2017). However, climate change poses a number of novel threats making it difficult to predict how these ecosystems will respond to marine protection. Warming oceans can have direct impacts on kelps via heat stress (Wernberg et al., 2016; Filbee-Dexter et al., 2020) and can also lead to the arrival of new organisms or the expansion of native species that alter ecosystem dynamics and often result in the loss of habitat forming kelps (Ling, 2008; Vergés et al., 2014). For example, the subtropical sea urchin Centrostephanus rodgersii (commonly known as the “long-spined sea urchin”) has extended its range into eastern Tasmania resulting in the extensive loss of kelp forests with flow-on impacts to biodiversity, ecosystem services and local fisheries (Johnson et al., 2005; Ling, 2008; Ling et al., 2009a; Perkins et al., 2015; Ling and Keane, 2018). While this range extension is related to long-term oceanic warming and the poleward expansion of the East Australian Current, it has also been facilitated by the overfishing of the spiny lobster Jasus edwardsii, which is considered the main predator of C. rodgersii in the region (Ling et al., 2009b; Johnson et al., 2011). The protection of predators, primarily J. edwardsii, in Tasmanian marine reserves has been shown to increase resilience to the initial formation of C. rodgersii barrens, but in new reserves where extensive barrens already occur, there was no evidence for the recovery of kelp forests (Perkins et al., 2020; Ling and Keane, 2021). This highlights how the abundance and impacts of range extenders can be exacerbated by the removal of predators through overfishing.

C. rodgersii also occurs along the northeastern coast of Aotearoa New Zealand where it is widely expected to increase in abundance and range in response to oceanic warming (Pecorino et al., 2013a; Thomas et al., 2021). Given the extent of habitat degradation seen in Tasmania (Johnson et al., 2005; Ling, 2008; Ling and Keane, 2018), there is concern over how a population increase would impact kelp forest communities in New Zealand. However, there is currently no information on the long-term population trends of C. rodgersii in this region and how it may respond to warming, overfishing of predators, or marine protection. While C. rodgersii has previously been described as having extended its range to northeastern New Zealand (Pecorino et al., 2012; Byrne and Andrew, 2020), it was first recorded in New Zealand in 1897 and is considered to be a native species (Fell, 1949). It is most common in areas that are strongly influenced by the warm East Auckland current, specifically the northern-most part of the country and at offshore islands along the northeast coast (Ayling, 1974; Riddell, 1980; Choat and Schiel, 1982; Cole et al., 1992; Pecorino et al., 2013a; Byrne and Andrew, 2020; Thomas et al., 2021).

Unlike in eastern Australia, C. rodgersii has historically only occurred in low numbers and is not known to form urchin barrens in New Zealand (Choat and Schiel, 1982; Schiel, 1984; Jones and Andrew, 1990; Shears and Babcock, 2004; Byrne and Andrew, 2020). Instead, the endemic sea urchin Evechinus chloroticus or “kina” is the dominant sea urchin species and forms extensive areas of urchin barrens on many rocky reefs in the region (Choat and Schiel, 1982; Andrew, 1988; Shears and Babcock, 2004). An increase in E. chloroticus abundance and prevalence of associated urchin barrens in northeastern New Zealand has been linked to overfishing of its major predators, snapper (Chrysophrys auratus) and spiny lobster (J. edwardsii), however within long-established marine reserves (30 years +), this pattern has been reversed with the recovery of predators and subsequent trophic interactions (Babcock et al., 1999; Shears and Babcock, 2002; Shears and Babcock, 2003). It is currently unknown whether these predators can play a similar role in controlling C. rodgersii populations within MPAs, and how these interactions may be influenced by climate change. While the rate of long-term ocean warming in northeastern New Zealand is lower than that in Tasmania (Shears and Bowen, 2017; Sutton and Bowen, 2019), the region has experienced large marine heatwaves in recent years (Salinger et al., 2019; Bell et al., 2022), and a more recent analysis of satellite-derived SST since 1982 suggests higher rates of warming in northeastern New Zealand than previously reported (Sutton and Bowen, 2019). Consequently, recent warming trends in the region may have promoted similar increases in C. rodgersii with wider consequences for rocky reef communities.

In this study, we examine long-term trends in both the endemic sea urchin E. chloroticus and the subtropical sea urchin C. rodgersii, and the extent of urchin barrens associated with both species, at a protected and unprotected offshore archipelago in northeastern, New Zealand: The Poor Knights Island Marine Reserve (PKI) and the Mokohinau Islands. The PKI have been fully protected from all forms of fishing since 1998, whereas the Mokohinau Is are heavily fished by recreational and commercial fishers. Predatory fish numbers, primarily C. auratus, have increased dramatically following protection at PKI (Denny et al., 2004; Schiel et al., 2018), but unlike mainland marine reserves there has been no evidence for the recovery of lobster (J. edwardsii and Sagmariasus verreauxi; Taylor et al., 2011). Shallow reefs at both locations were surveyed in 1999, 8 months after the PKI reserve was established, and found that while both sea urchin species were common, E. chloroticus was considerably more abundant and dominated the urchin barren habitat (Shears and Babcock, 2004; Shears et al., 2008). Ongoing monitoring of these sites from 1999-2022 provides a unique opportunity to examine how these two sea urchin species have responded to marine protection over a period of ocean warming. Based on long-term changes observed in coastal reserves, we expected the recovery of predators at PKI to result in a long-term decline in density of E. chloroticus and associated urchin barren habitat. In contrast, there is little understanding on how C. rodgersii may be impacted by marine protection or conversely how this species has responded to recent warming trends (Sutton and Bowen, 2019). Based on trends seen in Tasmania, we predicted a long-term increase in C. rodgersii in relation to warming, but this would be less evident in the reserve if predators were effective at controlling C. rodgersii densities and increasing resilience to barren formation. To better understand the potential drivers of changes over this period, we also provide an updated analysis of trends in satellite-derived SST data and quantified lobster populations at both locations in 2022 as there was limited available data on the abundance of lobster and their response to protection. Therefore, by comparing long-term trends in the two sea urchin species between a marine reserve and a fished location, we aimed to (1) gain a better understanding of long-term trends in C. rodgersii populations in northern New Zealand, and (2) assess the role of marine reserves as a management option for mitigating their impacts.

The densities and population structure of the two sea urchins, E. chloroticus and C. rodgersii, exhibited markedly different responses in relation marine reserve protection over the 23-year period. Densities of exposed E. chloroticus declined in the Poor Knights Islands Marine Reserve and urchin barrens associated with this species disappeared over the 23-year period. In contrast, densities of exposed E. chloroticus remained relatively stable at the fished location, with no long-term change in the extent of associated barrens. These results are consistent with a top-down effect of the very high abundance of predatory fish in the reserve (Denny et al., 2004; Schiel et al., 2018; Allard, 2020) and demonstrate how urchin barrens associated with this species can be reversed following marine protection. In contrast, the densities of the subtropical sea urchin C. rodgersii increased significantly at both reserve and non-reserve sites, with a concomitant emergence and increase in the extent of urchin barrens associated with this species ( Figure 1D ). The increase in this species coincides with significant ocean warming in northeastern New Zealand over the study period and the increase in the marine reserve suggests a lack of predators capable of preventing this outbreak. Despite full protection from fishing at the PKI, our surveys found very low numbers of rock lobster (J. edwardsii and S. verreauxi), which are likely the most important predators of C. rodgersii in northern New Zealand. Our results demonstrate the emerging threat C. rodgersii poses to rocky reefs in the region and how marine protection alone may not increase resilience to this threat.

Long-term population trends of the common sea urchin E. chloroticus followed similar patterns to those seen inside and outside other marine reserves in northeastern New Zealand, with long-term declines in exposed urchins and associated barrens within marine reserves due to increased predation (Babcock et al., 1999; Shears and Babcock, 2003; Edgar et al., 2017). The long-term decline in E. chloroticus and a shift to predominantly cryptic behaviour within the PKI marine reserve is most likely attributable to the large increase in abundance and size of C. auratus (Denny et al., 2004; Schiel et al., 2018; Allard, 2020), which are important predators of E. chloroticus (Shears and Babcock, 2002; Spyksma et al., 2017). Despite the rapid increase in C. auratus populations in the PKI marine reserve (Denny et al., 2004; Schiel et al., 2018; Allard, 2020), the declines in E. chloroticus numbers and associated barrens were gradual, taking ~20 years for a complete decline in E. chloroticus barren habitat to be seen. This is consistent with other marine reserves globally, where trophic interactions on non-fished species take on average 13 years to manifest, while fished species generally recover much faster (~5 years; Babcock et al., 2010). At the Mokohinau Is, where C. auratus are heavily targeted and remain at low levels (Denny et al., 2004; Allard, 2020), E. chloroticus densities and the cover of urchin barrens have generally remained stable. However, there was a sharp decrease in exposed urchins and concurrent increase in cryptic urchins between 2019 and 2022 at the Mokohinau Islands. This likely reflects a temporary shift in behaviour from exposed to cryptic caused by an extended bloom of the tropical dinoflagellate Ostreopsis siamensis that occurred at the Mokohinau Islands at the time of the 2022 survey (December 2021-May 2022; authors pers. obs.). O. siamensis has a number of impacts on E. chloroticus, including causing urchins to seek shelter and temporarily exhibit cryptic behaviour (Shears and Ross, 2009). The decline in the extent of E. chloroticus barrens at the Mokohinau Is between 2019 and 2022 may be also related to O. siamensis blooms as well as commercial harvest of E. chloroticus from some of the sites (~80 tonnes of E. chloroticus were harvested from the Mokohinau Is between 2017-2022, P. Herbert – Sea Urchin New Zealand, pers. comm.). Nevertheless, the general contrasting long-term patterns and trajectories between the marine reserve and fished locations provide further evidence on how long-term protection of predators can exert a top-down control on some sea urchin species and decrease the prevalence of urchin barrens.

In contrast, there was no apparent effect of marine reserve protection on C. rodgersii, with the population increasing steadily between 1999 and 2022 and the subsequent development of C. rodgersii barrens at both reserve and fished locations. This demonstrates that while the high population abundance of C. auratus at the PKI has been sufficient to control E. chloroticus populations, it has not had a regulatory effect on C. rodgersii populations. While C. auratus are known to be important predators of E. chloroticus, they are not known to predate on C. rodgersii which have considerably longer spines and are predominantly cryptic during the day (Ling et al., 2009b; Byrne and Andrew, 2020). Large rock lobster (J. edwardsii and S. verreauxi), on the other hand, can effectively predate on C. rodgersii (Ling et al., 2009b; Ling and Johnson, 2012; Provost et al., 2017; Day et al., 2021) and healthy populations of lobster have been demonstrated to regulate sea urchin populations and prevent the initial formation of C. rodgersii barrens in Tasmania (Ling et al., 2009b; Ling and Johnson, 2012). Rock lobster were very rare at both locations in this study, which is congruous with J. edwardsii being considered ecologically extinct in the wider region (MacDiarmid et al., 2013). The low densities (<1 lobster 500 m²) and lack of recovery of lobster following marine reserve protection is consistent with previous studies at PKI (MacDiarmid and Breen, 1992; Taylor et al., 2011). This strongly contrasts mainland marine reserves in New Zealand where lobster have increased following protection and occur at much greater abundances (>10 m²) than seen in the Poor Knights Islands marine reserve (Freeman et al., 2012). The reasons for the lack of recovery following protection from fishing at the PKI are generally unknown, but may include limited larval supply associated with the small and isolated islands, limited shallow reef habitat for settlement, and a low regional pool of larvae associated with a severely depleted spawning population (MacDiarmid and Breen, 1992; MacDiarmid et al., 2016; Hanns, 2021). Despite full no-take protection at the PKI, an absence of key predators and lack of trophic redundancy in both locations has meant that with warmer conditions more suitable for C. rodgersii, the populations have been able to increase unchecked.

This study provides the first documentation of an increase in C. rodgersii, and emergence of urchin barren habitat associated with this species in New Zealand. Only E. chloroticus has previously been reported as forming urchin barrens in the region (Choat and Schiel, 1982; Shears and Babcock, 2004) and earlier research at these study locations only report low densities of C. rodgersii with limited grazing impacts (Riddell, 1980; Schiel, 1984; Battershill, 1986; Berben et al., 1988; Cole, 1993). In our study, E. chloroticus barrens were common at both locations in 1999, and C. rodgersii was only present in a low number of quadrats (10%) and did not form barrens. Now, 23 years later, C. rodgersii are commonly recorded in quadrats (40%) and barrens associated with this species are a common feature of reef ecosystems at both the PKI and Mokohinau Islands. While quantitative data is lacking from other locations, there are anecdotal reports of similar increases in C. rodgersii over the last decade at other offshore islands in northeastern New Zealand and in areas of the mainland coast north of the PKI (authors pers. obs.). At both locations examined in the current study, C. rodgersii and associated barrens were generally found deeper than E. chloroticus barrens (>10 m), often with an area of overlap (“mixed barrens”) at intermediate depths (~7-10 m). The presence of barrens did not vary across the depths sampled at the Mokohinau Islands, but there was an increasing likelihood of barrens with depth at the PKI which likely related to the decline in E. chloroticus barrens in shallow and emergence of deeper C. rodgersii barrens. C. rodgersii were most abundant in the deepest depth substratum (10-12 m) at both locations, but also extended well beyond the depths monitored and has formed extensive barrens to depths of ~25 m at some monitoring sites (authors, unpublished. data). This deeper distribution has been observed in Tasmania with C. rodgersii forming barrens to depths of ~ 58 m (Perkins et al., 2015). Consequently, we have likely underestimated the overall extent of C. rodgersii barrens in each location. The greater depth range of C. rodgersii poses a threat to deeper kelp forests and other reef habitats that are not typically impacted by E. chloroticus grazing, suggesting it may pose a greater threat to reef biodiversity than E. chloroticus. The range expansion of C. rodgersii into Tasmania has been described as the “single largest biologically mediated threat to the integrity of important shallow water rocky reef communities” (Johnson et al., 2011), with extensive economic implications to environment, tourism and fisheries (Johnson et al., 2005; Ling and Keane, 2018). While these effects in New Zealand are likely in an earlier stage than Tasmania and their extent largely undocumented, C. rodgersii likely presents an increasing and serious threat to rocky reef ecosystems.

The increase in C. rodgersii observed in this study coincides with a period of warming ocean temperatures, which is consistent with trends observed in Tasmania (Johnson et al., 2005; Ling et al., 2008; Ling et al., 2009a) and earlier predictions for this species (Pecorino et al., 2013a; Byrne and Andrew, 2020). Recent evidence suggests local recruitment is playing an increasingly important role for New Zealand populations (Thomas et al., 2021) and our results clearly demonstrate a widening size structure at both locations over time. In 1999, only large individuals (≥80 mm TD) were recorded, which is consistent with earlier studies that noted an absence of juvenile C. rodgersii on northeastern New Zealand reefs (Cole, 1993). The large increase in density at PKI in 2022 coincided with higher numbers of small urchins, which may indicate higher recruitment associated with recent warm years. In eastern Tasmania, warmer ocean temperatures were a major contributing factor to the population expansion of C. rodgersii as this provided more favourable developmental conditions for larvae (Ling, 2008; Ling et al., 2009a). Previous research from northeastern New Zealand has shown that winter temperatures when C. rodgersii spawn were above the threshold for successful reproduction and recruitment albeit at the lower end of the temperature limit (Pecorino et al., 2013a; Pecorino et al., 2013b). Despite a warming trend during summer and autumn, there was no evidence of long-term warming in winter and spring which would include the time of spawning and larval dispersal. Nevertheless, recent marine heatwaves and consistently warmer than average conditions since 2015 across all months may have provided more favourable conditions for reproduction and larval development which is reflected in the emergence of juvenile C. rodgersii. While the exact mechanisms linking warming and population increases in northern New Zealand are unknown, it is likely that C. rodgersii populations will continue to increase in density with continued warming and increasingly self-sustaining populations (Thomas et al., 2021).

MPAs are traditionally considered an effective tool for rebuilding predator populations, and subsequently restoring trophic interactions and increasing ecosystem resilience (Roberts et al., 2017; Sala and Giakoumi, 2018; Eger and Baum, 2020; Peleg et al., 2023). However, as seen in this study, marine protection alone may not always restore the full complement of predators required to prevent irruptions of native species. While there is some contention over the role of lobster in controlling C. rodgersii populations and associated barren habitat (Glasby and Gibson, 2020; Smith et al., 2022; Kingsford and Byrne, 2023), rock lobster are likely the most important predator of C. rodgersii in New Zealand waters and they have not recovered in the marine reserve examined. In both Tasmania and now New Zealand, the increase in C. rodgersii coincides with ocean warming, but also follows a long history of exploitation of J. edwardsii, which has provided an opportunity for this species to proliferate, unconstrained from predatory control (Ling et al., 2009b; Ling et al., 2015). While MPAs with healthy lobster populations may be more resilient to future outbreaks of C. rodgersii (Ling et al., 2009b; Perkins et al., 2020; Ling and Keane, 2021), recovery of predators in MPAs may be limited by both local and regional factors. Furthermore, once extensive urchin barrens have established, these areas are extremely difficult to reverse and may take decades to recover, even with full marine protection (Shears and Babcock, 2003; Babcock et al., 2010; Ling et al., 2015; Perkins et al., 2020; Ling and Keane, 2021).

Consequently, there is an urgent need for an active and multi-faceted approach to controlling C. rodgersii in northern New Zealand in order to reduce their potential socio-economic and ecological impacts and build ecosystem resilience to future warming. While more research into the role of key predators such as rock lobster in controlling C. rodgersii populations is needed, rebuilding populations of key predators at large spatial scales is crucial to increase resilience and can be achieved by increasing the area of MPAs combined with large-scale reductions in fisheries catch. In areas where C. rodgersii have already established, active restoration through culling or harvest, or the introduction of lobsters, may also be needed to promote recovery (Tracey et al., 2015; Ling and Keane, 2021; Miller et al., 2022).

Global research into the effects of climate change on marine ecosystems has largely focused on how range extending species can lead to large-scale changes in ecosystems (Pecl et al., 2017). Our study demonstrates an irruption of a native species with the scope to have comparable ecosystem level consequences to that of range extending species. As seen in Tasmania, it is likely that this increase has been facilitated by a warming climate and the historic and current overharvesting of the only known predator in the region. Increases in native species can pose additional challenges to conservation and management (Carey et al., 2012), and in this study, marine protection did not provide the expected resilience to this irruption. Consequently, the absence of large predators from many marine ecosystems ultimately reduces ecosystem resilience and provides unabated opportunities for new species to take hold under changing environmental conditions. The rate and scale of the emergence and outbreaks of these species is therefore likely to intensify with climate change unless urgent and radical changes are made to how we manage marine ecosystems and fisheries (Lotze et al., 2011; Duarte et al., 2020).

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

NS designed and led the monitoring and data collection. CB performed the data analysis and wrote the manuscript with input from NS. CB and NS contributed to the manuscript revisions and approved the final version. All authors contributed to the article and approved the submitted version.