A review of urchin barrens and the longspined sea urchin (Centrostephanus rodgersii) in New South Wales, Australia

Introduction

Urchin barrens and the complex ecological interactions associated with them continue to be a topic of interest and debate among academic, industry, Indigenous, conservation and government groups (Andrew 2022; Environment and Communications References Committee 2023; Kingsford and Byrne 2023). In Australasia, much of this interest has focussed on the climate change-related range extension of the longspined sea urchin (Centrostephanus rodgersii) into Tasmania where it has driven an ecosystem shift from kelp forests to barrens (Johnson et al. 2005; Ling 2008; Ling and Keane 2024), as well as population increases in Victoria and New Zealand (Gorfine et al. 2012; Balemi and Shears 2023). Centrostephanus rodgersii is native to New South Wales (NSW) (Thomas et al. 2021), and its latitudinal distribution extends from far south Queensland to Wilson Promontory in Victoria (Andrew and Underwood 1989; Andrew 1993; Byrne and Andrew 2020), with an associated wide thermal tolerance (Table 1). There is concern from the public that the area of barrens habitat along the NSW south coast has recently increased whereas kelp cover has decreased, similar to changes observed in Victoria and Tasmania; these concerns were reflected in a 2023 Senate Enquiry (Environment and Communications References Committee 2023).

Table 1.Temperatures at the northern (warm) range edge and main northern distribution of Centrostephanus rodgersii, Caloundra, Qld, and Solitary Islands.

Location		Mean minimum (°C)	Mean (°C)	Mean maximum (°C)	Data source
Caloundra (warm range edge)		19.6 ± 0.01 (n = 10,151)	23.1 ± 0.01 (n = 112,700)	26.5 ± 0.01 (n = 7843)	Caloundra: 2013–21 (see https://www.data.qld.gov.au/dataset/coastal-data-system-waves-caloundra)
Solitary Islands and Coffs Harbour (warm range edge main distribution)		19.2 ± 0.01 (n = 11,904)	21.6 ± 0.01 (n = 129,409)	24.3 ± 0.01 (n = 10,417)	2001–10 Australian Institute of Marine Science (AIMS). In situ logger data, South Solitary Island, NSW, Jan 2001–Mar 2011 (see https://apps.aims.gov.au/metadata/view/a8149e3d-21eb-40ba-9d4c-8d19fba84beb, accessed 28 June 2021)
Jervis Bay (mid-range)		14.6 ± 0.01 (n = 8575)	18.4 ± 0.003 (n = 101,678)	22.4 ± 0.01 (n = 8064)	Jervis Bay in situ loggers 16 Sep–18 Nov (NSW DPI Fisheries, unpubl. data)
South-eastern Tasmania (cool range edge and edge main distribution, C. rodgersii)		11.8 ± 0.24	14.4 ± 1.3	17.6 ± 0.39	Modis-Aqua (see https://oceancolor.gsfc.nasa.gov/about/missions/aqua/)

Jervis Bay NSW approximates the mid-range of C. rodgersii, and its southern limit is south-eastern Tasmania (Byrne et al. 2022). Map indicates locations referred to throughout this review, with underlay from Google Earth.

Barrens associated with C. rodgersii are recognised as a typical and distinctive habitat of NSW rocky reef ecosystems (Underwood et al. 1991; Wright et al. 1997). Barrens are rocky reef areas covered by crustose coralline algae that can support distinct fauna (sponges, ascidians, urchins, limpets, fishes) (Andrew and Underwood 1989; Curley et al. 2002). Sea urchins are conspicuous in barrens and, in NSW, species include Centrostephanus rodgersii, Heliocidaris erythrogramma, Heliocidaris tuberculata, Pseudoboletia Indiana, Tripneustes australiae and Phyllacanthus parvispinus. In NSW, C. rodgersii is the dominant urchin species in barrens (Andrew and Underwood 1989; Connell and Irving 2008) and the main species responsible for grazing kelp, thus creating and maintaining this habitat (Fletcher 1987; Andrew 1993).

Subtidal rocky reefs in NSW are characterised by mosaics of barrens, turfing algae and macroalgae beds (Underwood et al. 1991) (Fig. 1), and urchin density is positively correlated with area of barrens (Davis et al. 2020). Outside NSW, barrens have been associated with low macro-biodiversity, low primary productivity (Ling 2008) and loss of blue carbon (Carnell et al. 2020). In these areas, barrens are often considered an undesired alternate state of kelp ecosystems (Filbee-Dexter and Scheibling 2014; Ling et al. 2015; Young et al. 2023) and, at a global scale, they have been identified lower-ranked than kelp forests in 11 of 15 ecosystem properties (Eger et al. 2024). By contrast, studies in NSW show high fish biodiversity in barrens (Curley et al. 2002) and comparable microscopic biodiversity in both systems (Coleman and Kennelly 2019), similar to what has been found in the Mediterranean with high abundance and biodiversity of benthic invertebrates in barrens (Agnetta et al. 2024).

Fig. 1.

Rocky reef at North Rock, Broughton Island, NSW, Australia (aerial image: 23/09/2014) with associated habitats: (a) barrens, (b) macroalgae, (c) sand and (d) turfing algae.

Centrostephanus rodgersii is a conspicuous species along the NSW coast, homing to crevices in rocky reefs during the day (Andrew 1993; Flukes et al. 2012) and moving out to feed at night (Jones and Andrew 1990; Smith et al. 2024). Centrostephanus rodgersii individuals seem unable to directionally sense macroalgal food and instead forage in a generalised pattern around their nocturnal home crevices (Flukes et al. 2012). Gut content analyses on NSW urchins shows a diverse diet of predominantly molluscs, crustaceans and potentially drift algae (Day et al. 2024a), and the species can also graze kelp gametophytes (Veenhof et al. 2023). Spawning is highly synchronous and occurs in winter (King et al. 1994), with southern populations having a longer spawning period than northern populations (Byrne et al. 1998). Although booms and busts in populations are recorded (Andrew 1991), most NSW populations have been shown to be stable over multiple years (Andrew and Underwood 1989; Glasby and Gibson 2020; McLaren et al. 2024). The oldest NSW individual was 27 years old on the basis of sclerochronological aging estimates from jaw growth ridges (Blount et al. 2024a), whereas in Tasmania the oldest individuals were estimated to be more than 50 years old on the basis of jaw growth (Ling and Johnson 2009). It is not clear if the different methods used in these studies are comparable. Size frequencies among sites are variable (Andrew and O’Neill 2000), although C. rodgersii individuals are smaller in barrens than nearshore mosaic habitats, and size structure shifts to smaller individuals from north to south (Blount et al. 2024b). Genetic diversity among populations is associated with sea surface temperature (SST) and geography rather than spatial proximity (Banks et al. 2007), likely reflecting the long larval duration (1–4 months depending on temperature) and the high dispersal potential of this species (Huggett et al. 2005; Mos et al. 2020).

Increases in urchin abundance have been linked to a reduction in the size of predators, possibly driven by fishing pressure (Andrew and Byrne 2006; Ling et al. 2009). The main predators of C. rodgersii in NSW are uncertain, but are likely to include the eastern rock lobster (Sagmariasus verreauxi) (Day et al. 2021), a generalist that can eat up to three urchins per day in captivity (Provost et al. 2017), and larger individuals of species such as snapper (Pagrus auratus) and eastern blue groper (Achoerodus viridis) (Gillanders 1995; Byrne and Andrew 2020). Previous and recent observations point to the horned sharks as potential predators of C. rodgersii in NSW (McLaughlin and O’Gower 1971; Day et al. 2024b). Abiotic stressors may also decrease urchin abundance, including ocean acidification, which could result in smaller larvae (Foo et al. 2012) and flooding events that can result in mass mortality (Andrew 1991; Webb and Byrne 2025). The larvae of C. rodgersii are broadly thermotolerant and may be comparatively resilient to concurrent warming and acidification because of the presence of tolerant genotypes (Foo et al. 2012; Byrne et al. 2022).

Centrostephanus rodgersii is the main species of urchin harvested in NSW, with annual commercial catches generally increasing over the past 20 years, ranging from 24 to 250 tonnes (Mg) per year (Chick 2020, 2023). Most harvest occurs from March to early May, when the gonads (roe) are at their maximum weight and suitable for commercial harvest (Byrne et al. 1998). Commercial harvest in NSW is managed through the NSW Sea Urchin and Turban Shell (SUTS) restricted fishery. The greatest yield of high-quality roe is collected during summer and autumn from moderately sized urchins (Andrew et al. 1998; Blount and Worthington 2002).

Previous reviews and data syntheses on the ecology of barrens and associated urchin densities have adopted a global scale (Filbee-Dexter and Scheibling 2014; Ling et al. 2015) or within regions outside the main distribution centre of C. rodgersii (Ling and Keane 2024). The aim of this review is to summarise the spatiotemporal patterns and potential impacts of C. rodgersii and barrens on the environment and fisheries of NSW, the centre of historic distribution of this species. In addition, we identify key knowledge gaps and provide suggestions for future research. This paper is framed according to questions commonly asked by stakeholders, including recreational, commercial and cultural fishers (Przeslawski et al. 2023). We anticipate that this review will inform more accurate, broader discussions about urchin barrens in eastern Australia and contribute to improved further inter-jurisdictional management and research.

Ecological importance: why do barrens and C. rodgersii matter?

It has been suggested that barrens support lower biodiversity than do kelp beds and other rocky reef habitat (Filbee-Dexter and Scheibling 2014; Medrano et al. 2020). However, this has not been quantitatively assessed in NSW except for abalone (Andrew et al. 1998). Indeed, barrens around Terrigal and Sydney in NSW typically support a greater number of fish species than do kelp beds and a similar number of species to sponge habitat (Curley et al. 2002). Densities of C. rodgersii can be negatively correlated with abalone densities (Andrew and Underwood 1992), thereby decreasing local productivity (Andrew and O’Neill 2000).

Barrens have also been linked to negative commercial impacts, including challenges harvesting abalone (Strain et al. 2013) and the quantity and quality of commercial urchin roe (Blount and Worthington 2002). The reproductive output of C. rodgersii in barrens can be lower than those living on the fringe of macroalgal beds, and urchins from barrens are generally not considered worth harvesting (Byrne et al. 1998). Urchins transplanted to habitat with an abundance of macro-algae showed significant improvements in colour and yield of roe after 6 weeks, with the magnitude of change greatest at low urchin densities and between October and January (Blount et al. 2017). Furthermore, significant improvements in both roe colour and yield from urchins in barrens occurred after urchin density was reduced by 66% over 3 months, with greater improvement after 2 years (Blount et al. 2017). However, smaller-scale experiments along similar sections of coastline found that indices of urchin gonad quality were similar between macroalgae and barrens habitat at four of five locations along the southern NSW coast (Day et al. 2024a). These results indicate strong spatiotemporal variation in roe quality of urchin within and among different habitats, including nearshore barrens. Urchin density and food availability are important drivers, and food subsidies provided by drift macroalgae also need to be considered.

Centrostephanus rodgersii is an ecologically important species along the NSW coast, playing a key role in the maintenance of the habitat mosaics of macroalgae, turfing algae and barrens that comprise NSW shallow rocky reefs (Andrew 1993) (Fig. 1). The species also acts as a biogenic habitat, with 80 taxa recorded from underneath 180 urchins along the central and southern NSW coast, predominantly gastropods, malacostracans and chitons (Davis 2002). Similarly, Montgomery (2023) found C. rodgersii was associated with a diverse guild of associates including amphipods, reef fish (Lepidoblennius haplodactylos, Aspasmogaster costata), decapods (Rhynchocinetes serratus, Pagurus lacertosus), gastropods (Astraea tentoriiformis, Turbo torquatus) and other echinoderms (Clarkcoma pulchra, Phyllacanthus parvispinus, Ptilometa australis). Although the abundance of some of these groups differ between macroalgal habitats and barrens, total richness and abundance does not (Davis 2002).

Spatial patterns: where do barrens and C. rodgersii occur in NSW?

There are 10 published NSW statewide or large-scale assessments of barrens or associated urchin densities, including 3 that used aerial imagery informed by ground truthing (Andrew and O’Neill 2000; Glasby et al. 2017; Glasby and Gibson 2020), 4 that used underwater imagery (Jordan et al. 2010; Perkins et al. 2015; Davis et al. 2020; Williams et al. 2020) and 3 that used diver transects (Underwood et al. 1991; Connell and Irving 2008; Blount et al. 2024b) (Table S1 of the Supplementary material). In addition, Reef Life Survey maintains an extensive database on urchin density, extending along the entire NSW coast (see reeflifesurvey.com/explorer/map). These data can be used to identify the distribution of urchins and how barrens may change over time.

Barrens occur across much of the NSW coastline where there are rocky reefs. There is a general trend for the extent of barrens in NSW to increase from north towards the south, although they are absent or cover small areas at various locations throughout the state (Underwood et al. 1991; Andrew and O’Neill 2000; Davis et al. 2020; Glasby and Gibson 2020). The largest patches of barrens occur on the NSW southern coast (Connell and Irving 2008) where they are estimated to cover >50% of nearshore rocky reefs (Underwood et al. 1991; Andrew and O’Neill 2000; Davis et al. 2020), whereas in northern NSW barrens generally cover <20% of rocky reefs (Underwood et al. 1991; Davis et al. 2020). Barrens do not often occur on rocky reefs that are affected by sand movement, such as the narrow sloped reefs in northern NSW (north of Port Stephens) and in the southern-most part of the state (south of Wonboyn) (Andrew and O’Neill 2000; Glasby and Gibson 2020) (Table 1), possibly because of scouring effects on urchin larvae or food sources.

The distribution of urchins and associated barrens in NSW is depth-related, with barrens restricted to waters >2 m, likely because of strong wave action in the surge zone that limits foraging of C. rodgersii (Davis 2002). Barrens occur most frequently at depths of <20 m (Jordan et al. 2010), but at some sites can be common in depths up to 30 m (e.g. cod grounds in Jordan et al. 2010). Interestingly, barrens do not seem to extend much deeper than 30 m, with Perkins et al. (2015) finding a lower depth limit of 27 m for NSW barrens despite surveying rocky reef down to 50 m.

The distribution and abundance of C. rodgersii in NSW does not have consistent spatial patterns among localities (Andrew and Underwood 1992), although the far-northern NSW has lower urchin densities reflecting the distribution limits of this species (Connell and Irving 2008; Davis et al. 2023). Densities of C. rodgersii in NSW show high spatio-temporal variation at some sites across seasons (Underwood et al. 1991), and average urchin densities range from <0.01 urchins per square metre around Tweed Heads and Cape Byron in northern NSW to 1.2 urchins per square metre around Eden along the far southern coast (Davis et al. 2020) (Table 1). The highest published densities of C. rodgersii recorded in New South Wales are ~24 urchins per square metre (fringing habitat in summer at Guerilla Point in 1988, Underwood et al. 1991; unspecified rocky reef habitat in autumn at Jervis Bay in 2009, Davis et al. 2023). This is more than the maximum density of C. rodgersii recorded in Tasmania at 5.4 urchins per square metre (fringing habitat in southern Tasmania in 2021) (John Keane, pers. comm.).

Temporal patterns: are the area of barrens or densities of C. rodgersii increasing in NSW?

The shallow subtidal reefs of NSW, including barrens, have been characterised and monitored since the 1980s, and aerial imagery is available in some locations since the 1960s (Andrew and Underwood 1989; Andrew and O’Neill 2000; Glasby and Gibson 2020) (Supplementary Table S1). In general, data over a 50-year time period indicate that barrens and C. rodgersii are a dominant yet stable feature of NSW shallow subtidal ecosystems (Hill et al. 2003; Glasby and Gibson 2020; McLaren et al. 2024).

There is evidence of natural variation (increases and decreases) in the area of barrens at large spatial scales (one to tens of kilometres) from the 1980s to 2010s, with most sites between Newcastle and Eden fluctuating by ±10% in area (Glasby and Gibson 2020). The greatest increase in area of barrens has occurred in the Sydney region, in which three sites had increases in barrens of 20–35% (Glasby and Gibson 2020). Small increases in the area of barrens of 10–15% occurred at Shellharbour and Bawley Point and ~20% at Bermagui, whereas other areas showed less than 5% decreases in barrens (Glasby and Gibson 2020). Averaged across the 21 sites that were sampled, the extent of barrens in NSW has increased at a rate of 0.2% per year between the 1980s and 2010s (Glasby and Gibson 2020). At a local scale, monitoring in the Batemans Marine Park on the southern coast indicates that barrens have remained stable or even decreased in their coverage from 2008 to 2012 (Coleman et al. 2013). However, there is anecdotal evidence that kelp patches at particular sites in southern NSW have been replaced by barrens in the 2000s (Connell and Irving 2008). It is possible that different perceptions about changes in barrens in NSW could relate to the spatial scale of observation. Many of the aforementioned studies used coarse resolution (tens to hundreds of square metres) to track changes in the aerial extent of barrens in shallow waters, the size of which can be considerable (Connell and Irving 2008). Apparent stability of barrens at large scales (e.g. over hundreds of square metres) can mask fluctuations in kelp patches at smaller scales, noting that these could be either permanent or short-term losses.

Examples of longer-term persistence of extensive barrens are rare because underwater surveys began only in about the 1980s, but some large barrens in NSW have persisted for over 30 years and, in one case where longer-term data were available, up to 68 years (Glasby and Gibson 2020). Quantitative data are not available to assess whether the abundances of urchins and the extent of barrens areas have been increasing over long temporal scales (>50 years). However, recent analysis of yarning circles with Yuin Traditional Owners along the NSW southern coast suggests that barrens may have increased since European colonisation, which knowledge holders attribute to fewer abalone and lobsters (Chewying et al., in press). Thus, the comparatively recent research on NSW barrens may reflect shifting baselines, in which people perceive the state of environment as normal at a given recent time, thereby making it harder to detect environmental degradation across intergenerational timeframes (e.g. tens or hundreds of years) (Atmore et al. 2021). Importantly, large-scale ecological changes since colonisation have occurred across many Australian marine ecosystems (Daley 2014; Gillies et al. 2018; Statton et al. 2018), and a return to pre-colonisation conditions is almost certainly impossible, especially with a changing climate.

It is well documented that C. rodgersii has extended its range south to Tasmania since the turn of the century (Johnson et al. 2005; Ling and Keane 2024), and there is evidence suggesting that abundance may be increasing along the far-southern NSW coast (Davis et al. 2023) (Fig. 2a, b). In contrast to most of the state, the far-southern coast of NSW may be experiencing increases in urchin densities (Davis et al. 2023), and this may result in larger and more numerous barrens (Davis et al. 2020). Conversely, densities of C. rodgersii may have already declined at the distribution limits in northern New South Wales (Fig. 2a, b). However, where C. rodgersii co-occurs with the subtropical endemic coral Pocillopora aliciae in northern NSW, urchin densities have been stable over a 10-year period, including periods of significant heatwaves (McLaren et al. 2024). Through its grazing activity and associated reduction in coral competition with algae, C. rodgersii may be facilitating coral populations in the region, as was found for this urchin where it now co-occurs with corals in its extended range (Ling et al. 2018).

Fig. 2.

Centrostephanus rodgersii urchin density predictions and projections made using the optimal explanatory model developed for Australia. (a) Predictions for average densities from 1990 to 2000, (b) predictions for average densities from 2010 to 2020, (c) projections for average densities under climate change scenario RCP 8.5 for the period 2090–2100. Contours show average summer maximum (February) temperatures for the reference period. Adapted from Davis et al. (2023).

Climate change should be considered when considering local fluctuations in the extent of barrens and densities of urchins. Climate-related extreme events such as the eastern coast lows (extratropical cyclonic conditions in Dowdy et al. 2019) are a feature of the NSW coast. These storms can dislodge thousands of individuals weakened by low salinity (Byrne and Andrew 2020) and drive periodic mass mortality of C. rodgersii such as that observed in Botany Bay in 1986 and 1988 (Andrew 1991) and Botany Bay and Sydney Harbour in 2022 (Fig. 3). Although extreme storms and low salinity are also deleterious for kelp (Ebeling et al. 1985), it appears to recover quickly in NSW from flooding events (Davis et al. 2022a). Climate change will also involve warming waters, and predictions based on the business-as-usual emissions scenario (RCP 8.5) (Intergovernmental panel on Climate Change 2021) show continued range shifts of C. rodgersii such that populations are considerably reduced in all but the far southern coast of NSW (Davis et al. 2023) (Fig. 2c). These observations stress the importance of considering the impacts of changing climate and associated multiple stressors on the dynamics of NSW barrens and kelp habitats, including the response of thermally acclimated local populations of C. rodgersii in northern NSW and along the NSW latitudinal thermal gradient.

Fig. 3.

Mass mortalities of C. rodgersii and other urchins along Fairlight Beach in July 2022 after severe storms related to an eastern coast low. Image: Claire Reymond.

When considering temporal dynamics of barrens, kelp ecosystems are also important to understand as a potential alternate stable state (Filbee-Dexter and Scheibling 2014; Kriegisch et al. 2016). In NSW, kelp may be more vulnerable to stressors other than urchin herbivory. Mass mortalities and range contractions of Ecklonia radiata have been attributed to marine heat waves (Bosch et al. 2022), storms (Ettinger-Epstein and Kingsford 2008), disease (Valentine and Johnson 2004), poor water quality (Gorman et al. 2009) and flooding (Davis et al. 2022a). Similarly, declines in Phyllospora comosa have been linked to poor water quality (Coleman et al. 2008), warming and disease (Valentine and Johnson 2004). Although loss of kelp forests can occur quickly, natural recovery can take decades depending on the scale and previous history of the ecosystem. Furthermore, herbivorous fish can increase mortality and hinder recovery of kelp in both barrens and vegetated habitats (Andrew 1994), and herbivory is increasing in low latitudes where warmer waters are increasing herbivorous fish abundance (Vergés et al. 2016; Bosch et al. 2022).

Overall, there is a lack of evidence that C. rodgersii abundance or barrens area have increased along the NSW coast since the 1960s, but limited data preceding this time period make the identification of historical patterns challenging. It remains unclear whether the extent of barrens in NSW is a long-term natural phenomenon or a result of anthropogenic impacts (e.g. climate change, overfishing) that precede the ecological research cited here. Predictions based on the worst-case climate change scenario show that C. rodgersii may decrease in northern NSW and increase along the far NSW coast over the next 100 years (Davis et al. 2023), although local populations may be more resilient to warming waters (McLaren et al. 2024).

National and global context: how do NSW barrens compare to barrens elsewhere?

In NSW, scientists have not yet documented broadscale increases in barrens or obvious causes of barrens increase in local areas. This differs from our understanding of barrens ecosystems investigated elsewhere in the world. In many regions, barrens are the result of disturbances to rocky reef ecosystems (Ebeling et al. 1985; Eger et al. 2022). In the USA and Mediterranean, reductions in urchin predators through commercial harvest have enabled urchins to reach high densities, and they have overgrazed foliose algae on rocky reefs and substantially altered the rocky reef biodiversity (Guidetti and Dulčić 2007; Eurich et al. 2014). Importantly, the cyclical changes from kelp forests to barrens and the return to kelp forests on both sides of the North American continent are driven by sea urchin disease, periodic warming and mass mortality of predator species (seastars, sea otters) (Scheibling and Lauzon-Guay 2010; Rogers-Bennett and Catton 2019; McPherson et al. 2021; Smith et al. 2022a; Galloway et al. 2023). Such large-scale events have not yet been observed in Australia, and the drivers of barrens appear to differ from elsewhere in the world, possibly because of species-specific urchin behaviour, habitat complexity, geomorphology and oceanography (see discussion in Glasby and Gibson 2020).

Within Australia, unexplained short-term population booms of urchins periodically occur, along with associated increases in barrens, for example, Heliocidaris erythrogramma in NSW (Wright et al. 2005) and Victoria (Carnell and Keough 2019) and Tripneustes australiae at Lord Howe Island (Valentine and Edgar 2010; McLaren et al. 2023) and along the NSW coast (McLaren et al. 2024). These population booms may have contributed to the rapid increase of C. rodgersii in Tasmania after its initial range extension. This range extension and warming has threatened kelp (Ling 2008; Mabin et al. 2019) and is predicted to continue and increase the extent of barrens in Tasmania (Perkins et al. 2020). Recent estimates in local-scale studies (5–40-m depth) range from 0.018% barrens cover in southern Tasmania to 2.10% barrens cover in eastern Tasmania (Sward et al. 2022). Statewide estimates show over 40% continuous barrens at the reef at St Helens in north-eastern Tasmania and 15% averaged across the eastern coast (4–40 m) (Ling and Keane 2024). Barrens occur in significantly deeper waters in Tasmania (16–58 m) than in NSW (7–27 m) (Perkins et al. 2015). The incursion of warm low-nutrient EAC water, which will continue as a result of climate change, is a major threat to kelp in Tasmania (Mabin et al. 2019).

Management: how can barrens be managed in NSW?

Community and Indigenous groups are concerned by the presence of extensive barrens and large numbers of urchins on NSW rocky reefs (Chewying et al., in press), and there is growing interest in the idea of transforming barrens to kelp forests by harvesting or culling urchins (Blount 2022). The removal of all visible C. rodgersii individuals causes an ecosystem shift from barrens to habitats dominated by foliose algae if the abundance of urchins is kept very low for an extended period of time (Fletcher 1987; Andrew et al. 1998). Partial removal of even up to 66% of C. rodgersii in NSW barrens does not allow kelp and other foliose algae to successfully colonise (Andrew and Underwood 1993; Hill et al. 2003), with recent evidence indicating that this may be due to the remaining urchins compensating for the reduced abundance with faster growth rates and higher fecundity, which would be associated with increased herbivory. Although ineffective at transforming barrens, removal of some urchins can improve the quality of commercial product (roe) in the remaining animals (Blount et al. 2017). A global review found that removing sea urchins is unlikely to be a long-term solution to support kelp forests because it does not address the underlying cause of high urchin abundance (Miller et al. 2022). There has been research in Victoria and Tasmania showing varying degrees of success with C. rodgersii removal and kelp recovery (Gorfine et al. 2012; Sanderson et al. 2015; Tracey et al. 2015; Kriegisch et al. 2016), but it is challenging to compare these results to those in NSW because of different measurements (i.e. NSW surveys presented as percentage of original population reduction, whereas Tasmania surveys presented as urchin density) and ecosystems (i.e. C. rodgersii is native to NSW). Moreover, the range extension of C. rodgersii in Tasmania is driven by larval dispersal through the strengthening of the EAC (Johnson et al. 2005), an ecological construct different from the situation in NSW.

With the commencement of marine parks in NSW, it was initially predicted that the protection of urchin predators in no-take sanctuary zones may lead to a reduction in urchin numbers and associated barrens areas (Babcock et al. 1999; Ling and Johnson 2012). This prediction was based on findings from overseas, in which the restoration of predators within marine protected areas led to a shift from barrens to seaweed forests (Sangil et al. 2012; Edgar et al. 2017; Kawamata and Taino 2021), with similar patterns also found in New Zealand (Kerr et al. 2024). This is a field of research that is quickly developing, with improved understanding of predator–prey dynamics in NSW (Day et al. 2021, 2024b, 2024c) and Tasmania (Smith et al. 2022b, 2023), as well as the multi-decadal analysis of urchins, barrens and kelp habitat-mosaic in NSW (Glasby and Gibson 2020; Davis et al. 2022b, 2023).

In NSW, there is no evidence of reductions in barrens areas or urchin numbers in Marine Park Sanctuary Zones (Glasby and Gibson 2020), despite significant and widespread increases in the abundance in these no-take zones of the predators, C. auratus, Achoerodus viridis and potentially Sagmariasus verreauxi (Lee et al. 2015; Day 2020; Knott et al. 2021). Barrens are even more abundant inside of the sanctuary zones for Jervis Bay Marine Park than they are outside (Barrett et al. 2008). Research suggests that the control of C. rodgersii by lobsters has been overestimated, at least with contemporary lobster size distributions. Although S. verreauxi can eat up to three urchins per day when starved in aquaria (Provost et al. 2017), gut content analysis of this species suggests that urchins may not be a key food item, with C. rodgersii detected in only 1–2% of lobsters collected over a wide latitudinal range (Day et al. 2021, 2024b). This was corroborated by feeding trials, in which H. erythrogramma was eaten more regularly than C. rodgersii by S. verreauxi in NSW (Day et al. 2021), and small urchins of either species were more frequently consumed than were their larger counterparts (Day et al. 2024c). Moreover, because S. verreauxi undertakes extensive migrations (Booth 1997), it is not a permanent resident of C. rodgersii barrens.

Surprisingly, most predation of urchins occurred by small lobsters (<125-mm carapace length, CL) rather than larger individuals (Day et al. 2021, 2024c). Results of urchin tethering experiments in open barrens and fringe habitats (Day et al. 2023) and outside of an inhabited lobster den (Day et al. 2024b) support this result of low urchin predation by lobsters, as fish predators ate most tethered urchins irrespective of urchin size or species. Overall, these studies suggest lobsters as capable but hesitant C. rodgersii predators (Day et al. 2021). However, S. verreauxi is reportedly the largest growing lobster in the world (Montgomery and Liggins 2012) and research is ongoing into whether urchins including C. rodgersii are a significant food source for very large lobsters of >180 mm CL, whose populations have become reduced since European colonisation. Fish predators (including horned sharks) have more potential impact on urchin populations than do lobsters, because they ate the majority of urchins in tethering experiments (Day et al. 2023, 2024b, 2024c).

Overall, management of NSW barrens through the removal of C. rodgersii from subtidal reefs requires justification. Barrens formation and changes result from various mechanisms and vary over decades (Glasby and Gibson 2020). The expansion of barrens at a local scale, resulting in an undesirable outcome (e.g. a persistent loss of macroalgae), may support a management response (e.g. removal of urchins). To effect a change from barrens to macroalgal beds requires the removal of >70% urchins at a given site (Fletcher 1987; Andrew 1993; Hill et al. 2003). Removal efforts need to be sustained over time and are costly so the feasibility of long-term urchin removal and maintenance must be considered. Importantly, robust monitoring of the manipulated and control areas must occur to assess the effectiveness of urchin removals.

Future research needs: what are the knowledge gaps?

Owing to the different roles of C. rodgersii and urchin barrens between NSW and other parts of Australia, some research needs are unique to NSW whereas others are common to jurisdictions throughout this species’ range.

Conclusions

Barrens are a conspicuous part of the mosaic of habitats on NSW rocky reefs. Generally, the extent of barrens and densities of C. rodgersii in NSW increase from north to south. There is a lack of evidence that C. rodgersii abundance or barrens area have increased along the NSW coast since the 1960s. However, historical patterns are unable to be assessed, and shifting baselines may affect our understanding of barrens as a natural characteristic of NSW rocky reefs.

In conjunction with recent perspective pieces (Andrew 2022; Kingsford and Byrne 2023), this review has shown that barrens and the associated urchin C. rodgersii are ecologically important to NSW rocky ecosystems. In most areas across the state, they do not seem to be a problem to be managed as they are in other regions (Gorfine et al. 2012; Balemi and Shears 2023; Ling and Keane 2024). These findings can inform broader discussions at a national and international level to ensure that appropriate research, management, and policies are being implemented for each jurisdiction. Importantly, the temporal dynamics of any potential ecosystem shift from kelp forest to barrens will be influenced by climate-driven habitat warming because of the thermal sensitivity of kelp (Provost et al. 2017) and will be influenced by the frequency and severity of storms, with storms having the potential to clear large areas of kelp forest (Davis et al. 2022a). Without reducing these stressors or otherwise mitigating their effects, the removal of urchins and the restoration of kelp beds may yield few tangible benefits in the long-term.

Supplementary material

All published studies focussing on barrens or C. rodgersii in New South Wales as of June 2024 are available in Table S1 of the Supplementary material, which is available online.