Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

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Journal of Experimental Marine Biology and Ecology

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Regime shift of a kelp-forest benthic community induced by an ‘invasion’ of the rocklobster Jasus lalandii

Laura K. Blamey ⁎, George M. BranchMarine Research (MA-RE) Institute, University of Cape Town, Private Bag X3, Rondebosch, 7701, South Africa

⁎ Corresponding author. Tel.: +27 21 650 5454; fax:E-mail address: laura.k.blamey@gmail.com (L.K. Blam

0022-0981/$ – see front matter © 2012 Elsevier B.V. Alldoi:10.1016/j.jembe.2012.03.022

a b s t r a c t

a r t i c l e i n f o

Article history:Received 21 February 2012Received in revised form 22 March 2012Accepted 23 March 2012Available online xxxx

Keywords:BenthicKelp-forestLobsterRegime shiftTop-down controlUrchin

An eastward shift in the center of abundance of the South African rock lobster Jasus lalandii occurred duringthe early 1990s into an area known as East of Cape Hangklip (EOCH). Given (1) the predatory capabilities of J.lalandii, (2) an intricate relationship between the urchin Parechinus angulosus and juveniles of the abaloneHaliotis midae, and (3) existing over-exploitation of elements of the ecosystem, the ‘lobster invasion’ hasmajor implications for the benthic ecosystem and associated fisheries. We surveyed the abundance of J.lalandii and the benthic community composition EOCH at six sites (three ‘invaded’ and three ‘non-invaded’by J. lalandii), in three different depth zones (b5 m, 6–12 m and 13–20 m). At all depths, J. lalandii wassignificantly more abundant in invaded areas than in non-invaded areas, and benthic communities weresignificantly different. The high densities of rock lobsters at invaded sites led to cascading effects includingelimination of urchins and depletion of grazers, consequent enhancement of macroalgae, and diminishmentof encrusting corallines. Non-invaded sites had few lobsters, abundant herbivores, less macroalgae and moreencrusting corallines. Elimination of urchins in the invaded area has important implications, as juveniles ofthe commercially harvested abalone H. midae depend on shelter beneath urchins in this region. Floral speciesdiversity was greater at invaded sites and increased with depth, whereas faunal species diversity was greaterat non-invaded sites but also increased with depth. The depths at which strongest effects of J. lalandii werefelt coincided with the depth of maximum abundance of the urchin P. angulosus, the abalone H. midae, thekelp Ecklonia maxima and encrusting corallines, with serious consequences for associated fisheries and thebenthic community. We argue that the differences between invaded and non-invaded areas are sufficient torecognize that they are alternate stable-states and constitute a regime shift.

© 2012 Elsevier B.V. All rights reserved.

1. Introduction

Predators are capable of transforming the structure and func-tioning of ecosystems, to the extent that they may generatealternative stable states (Barkai and McQuaid, 1988). In this paperwe draw on previous experimental work and employ comparativestudies to document how the arrival of high densities of the rocklobster J. lalandii in an area where they were previously rare hasradically altered the community composition of the kelp beds theyoccupy. In doing so, we probe topical ecological concepts includingtop-down influences of predators, cascade effects, regime shifts andgeographic shifts related to climate change. We argue that thechanges we document are of sufficient magnitude to be regarded as a‘regime shift’ (Sensu Cury and Shannon, 2004; deYoung et al., 2004)and that the attending conditions and duration meet the require-ments for recognition of an alternative stable state, also known as a

+27 21 650 4930.ey).

rights reserved.

discontinuous regime shift (Collie et al., 2004; Connell and Sousa,1983; deYoung et al., 2004; Scheffer and Carpenter, 2003).

Kelp forest ecosystems are both diverse and highly productive, andmuch research has focused on trophic linkswithin them (see reviews inBranch, 2007; Estes, 2007; Fariña et al., 2007; Graham et al., 2007),particularly on algal–herbivore–predator interactions (Andrew, 1993;Byrnes et al., 2006; Dayton, 1985a; Estes and Duggins, 1995; Estes et al.,1998; Hagen, 1995; Halpern et al., 2006; Lang and Mann, 1976;Lawrence, 1975; Shears and Babcock, 2002; Tegner and Dayton, 2000;Watson and Estes, 2011).

Sea urchins in particular are integral to the functioning of kelp forestecosystems: because of their considerable grazing capabilities, they havebecome renowned for their potential to transform kelp forests intocoralline-dominated ‘urchin barrens’dominated by encrusting corallines(Andrew, 1993; Breen and Mann, 1976; Dayton et al., 1998; Estes andDuggins, 1995; Lawrence, 1975; Ling and Johnson, 2009; Mann, 1977;Steneck, 1998; Steneck et al., 2002; Tegner and Dayton, 1991;Watanabeand Harrold, 1991). Most examples of such kelp-deforestation are fromthe Northern Hemisphere, partly because several southern-hemispherespecies of urchins feed mainly on drift algae rather than on attachedplants (Castilla and Moreno, 1982; Day and Branch, 2002b; Dayton,

34 L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

1985b; Santelices and Ojeda, 1984; Vanderklift and Kendrick, 2005), sothey are incapable of inducing kelp deforestation (see review in Stenecket al., 2002). Nevertheless, denudation of kelp beds by urchins has beenrecorded in Australia (Ling, 2008; Ling and Johnson, 2009) and in NewZealand (Babcock et al., 1999).

The intense grazing pressure exerted by sea urchins on kelpcommunities is often controlled by the top-down effects of predators,including sea otters (Estes and Palmisano, 1974; Estes et al., 1989, 1998;Watson and Estes, 2011), fish (Cowen, 1983; Guidetti, 2006; Sala andZabala, 1996; Sala et al., 1998; Shears and Babcock, 2002, 2003; Stenecket al., 2004), clawed lobsters (Breen and Mann, 1976; Mann and Breen,1972) and rock (or spiny) lobsters (Babcock et al., 1999; Ling et al.,2009b; Mayfield and Branch, 2000; Shears and Babcock, 2002, 2003;Tegner and Dayton, 1981; Tegner and Levin, 1983).

Along the South African coastline, forests of the kelps Eckloniamaxima and Laminaria pallida are restricted to the cool-temperateshallow sub-tidal waters of the west and south-west coasts (Field etal., 1977; Velimirov et al., 1977), but important differences in kelp-forest community structure exist between these two coasts. Forestsalong the productive but low-diversity, upwelling-fuelled west coast(Shannon, 1985) have been recognized for their large numbers ofmussels and the West Coast rock lobster J. lalandii, and dense redalgae cover most of the substratum in the shallows (Anderson et al.,1997; Field et al., 1980a). In contrast, the south-west coast, which isless productive because upwelling is diminished there, supports ahigher number of species and endemics (Awad et al., 2002; Turpie etal., 2000) and has a greater abundance of herbivores such as the Capeurchin Parechinus angulosus, the abalone Haliotis midae and turbansnails Turbo sarmaticus and Turbo cidaris (Field et al., 1980a), butmussels are scarce there (Field et al., 1980b) and foliar algae are lessabundant, being replaced by extensive beds of encrusting corallines(Anderson et al., 1997).

As documented by Tarr et al. (1996), Mayfield and Branch(2000), and Blamey et al. (2010), J. lalandii shifted its center ofabundance south-eastward in the early 1990s, declining on the westcoast, and substantially increasing in kelp beds along the south-westcoast in an area known as ‘East of Cape Hangklip’ (EOCH), where

Invaded state

Jasus lalandii

Grazers

Macroalgae

Encrustingcorallines

Sessileanimals

Scavengers

S

-

Urchins

Predation

- -

-Grazing

--

Competitionfor space

-Competition

-Competition& Shading

PredationPredation

Predation

Fig. 1. A priori flow model summarizing hypothesized states and interactions of the ecosysteffects. Positive (+) and negative (−) effects are indicated in boxes. Sizes of circles indicat

their increased abundance has been termed an ‘invasion’. Reasonsfor this shift are largely unknown, but it is thought to be linked tochanges in environmental conditions. Roy et al. (2007) and Rouaultet al. (2009) have recorded a cooling of inshore waters on the southcoast since the 1980s, particularly east of Cape Agulhas, but long-term data are inadequate to test whether this could have accountedfor the expansion of the lobster population (Cockcroft et al., 2008).

The potential consequences of this invasion are immense, andextend to the economically important abalone, H. midae, which has anintricate relationship with the urchin P. angulosus (Day and Branch,2000a,b, 2002a). Unlikemany other urchins that act as grazers (Andrew,1993; Andrew and Underwood, 1993; Lawrence, 1975; Mann, 1977), P.angulosus feeds mainly by trapping drift kelp (Velimirov and Griffiths,1979; Velimirov et al., 1977), and remains relatively immobile, allowingjuvenile abalone (spanning 3–35 mm) to take refuge under its spinesand to secure nourishment from pieces of drift algae trapped by theurchins (Day and Branch, 2002a,b). The significant increase in J. lalandiiEOCH during the early 1990s coincided with a rapid decline and virtualdisappearance of P. angulosus, associated with a collapse in the densitiesof juvenile abalone (Blamey et al., 2010; Tarr et al., 1996). In that samedecade, abalone became a target for illegal fishing EOCH (Hauck andSweijd, 1999). This stretch of coastline once supported a major abalonefishery, but the combined effects of intense illegal fishing and urchindepletion associated with the rock lobster invasion have led to itsdemise (Hauck, 2009; Tarr et al., 1996). In its place, a fishery for J. lalandiihas been developed (DEAT, 2005). Several other resources involved inthe kelp-bed foodweb are also harvested, including linefish and kelp.

Benthic communities are likely to have been substantially altered bythe rock lobster ‘invasion’ and in this paper we explore the magnitudeand nature of these effects by comparing ‘invaded’ areas with adjacent‘non-invaded’ areas.

As a first step, we hypothesized a priori interaction webs for bothinvaded and non-invaded states, summarized in Fig. 1. In invadedareas where J. lalandii numbers are now high, we anticipated thatpredation by them would reduce scavengers, herbivores (urchins andgrazers) and sessile animals. The decline in herbivores was expectedto lead to a proliferation of macroalgae (foliar algae, turfs and kelps),

Non-invaded state

Jasus lalandii

Grazers

Encrustingcorallines

Sessileanimals

cavengers

Macroalgae

Urchins

-- -

-Grazing

-

--

+

Removal of competitive algae

+Settlement on

substrata?

Predation

PredationPredation

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Competitionfor space

Competition& Shading

ems in areas ‘invaded’ and ‘non-invaded’ by rock lobster. The bold arrows imply stronge relative biomass of different functional groups.

35L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

with indirect negative effects on sessile animals and encrusting corallinesthrough shading and competition for space. Encrusting corallines wereexpected to be additionally depleted by direct consumption by rocklobsters (Mayfield et al., 2000a,b).

The opposite circumstances were hypothesized for non-invadedareas where densities of J. lalandii are low. As a consequence oflimited predation by rock lobsters, we expected scavengers, herbi-vores and sessile animals to be more abundant, macroalgae to bedepleted and their negative effects on encrusting corallines to bereduced.

Against the backdrop of these hypothesized contrasting states, wefirst compared rock lobster densities among six sites – three putatively‘invaded’ and three ‘non-invaded’ – as a means of validating theirclassification in these two states. We then compared the composition ofbenthic communities between ‘invaded’ and ‘non-invaded’ sites todetermine the extent to which they conformed to the composition wehad hypothesized a priori.

2. Methods

2.1. Study sites

Data were collected from six study sites (Fig. 2) along the south-west coast of South Africa; all falling within the area termed ‘East ofCape Hangklip’ (EOCH). On the basis of fisheries surveys showing awest to east decline in rock lobster abundance and an abruptreduction in catches and catch rates east of Hermanus (Cockcroft etal., 2008), we regarded the three sites west of Hermanus as being‘invaded’ by high densities of rock lobsters, and those to the east as‘non-invaded’ with low densities. This a priori classification was thenverified by counts of rock lobsters and Fisheries IndependentMonitoring Surveys (see below). The six sites were chosen becausethey all (1) fell in the same biogeographic region, (2) experiencedcomparable moderately strong wave action, (3) supported kelp beds,and (4) had been sampled at least once in the past (with theexception of Quoin Point). The Betty's Bay site falls within a marineprotected area, in which removal of invertebrates is legally prohib-ited, although illegal harvesting of rock lobsters and abalone doestake place.

Fig. 2. Map showing the six (named) study sites along the south-west coast of South AfriIndependent Monitoring Surveys (FIMS).

2.2. Fisheries Independent Monitoring Surveys (FIMS)

Annual West Coast rock lobster inshore monitoring surveys wereconducted EOCH by the Department of Agriculture, Forestry andFisheries (DAFF) at set stations (Fig. 2) during Nov–Dec 2002–2005.At each station, 15 hoopnets were set in groups of five at three depthintervals: 0–10 m, 11–20 m and 21–30 m. Soak time was standardizedto 15–20 min for each hoopnet. The sex and carapace length (CL) ofeach lobster caught were recorded, and CPUE was defined as thenumber of lobsters caught per hoopnet per soak period. We employedthe data for Stations 61–70 (invaded) and Stations 71–90 (non-invaded), but excluded Stations 72–76 and 81–83 because they weredominated by sandy bottom unsuitable for rock lobsters.

2.3. Benthic community surveys

Sampling was carried out using SCUBA during 2005 and 2006.Three depth ranges were sampled: 2–3 m for the b5 m zone, 8–9 mfor the 6–12 m zone and 16–18 m for the 13–20 m zone. In eachdepth range, three 10-m transects were set parallel to the shore, atleast 5 m apart, on rocky substratum with a slope b30°. Along eachtransect six equidistant replicate 0.25 m2 quadrats were sampled, anda swim search conducted in which all rock lobsters, crabs, octopus,large abalone and giant periwinkles were counted within 1 m eitherside of the transect. Within each quadrat, algae and invertebrateswere identified to the level of species and quantified as percentagecover (sessile species) or counts (mobile species). Organisms thatcould not be identified in situ were collected for later identification.Sponges and compound ascidians were not identified beyond thelevel of class. The kelps E. maxima and L. pallidawere counted in 1 m2

quadrats if their stipe length exceeded 30 cm; smaller individualswere recorded as ‘juveniles’ in the 0.25 m2 quadrats.

2.4. Definition of functional groups

For some analyses, individual taxa were pooled into functionalgroups. J. lalandiiwas the dominant benthic predator and the functionalgroup ‘predator’ refers only to J. lalandii. All other benthic predatory/carnivorous species were grouped as ‘scavengers’. ‘Urchins’ constituted

ca, East of Cape Hangklip (EOCH), and the 31 (numbered) stations used for Fisheries

36 L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

the urchin P. angulosus, and ‘grazers’ comprised all other herbivores.‘Herbivores’ were defined as the combination of ‘urchins’ and ‘grazers’.‘Sessile species’ comprised attached faunal species, including mussels,barnacles, hydroids, sponges, ascidians and crinoids. The term ‘macro-algae’ incorporated all erect algae – turfs, foliar algae and kelps – asdistinct from encrusting corallines, which were treated as a separatefunctional group. A list of all taxa in these functional groups appears inBlamey (2010, Appendix A1.01).

2.5. Statistical analyses

Data on benthic communities obtained from the quadrats andtransects were converted from percentage cover or counts into biomass(grams wet weight per m2) either by weighing entire samples or byusing conversions based on subsamples (Appendix A1.01 in Blamey,2010).

2.5.1. Multivariate analysesFor each depth range, biomass data were analyzed using PRIMER

(Plymouth Routines in Multivariate Ecological Research, version6.1.5) (Clarke and Gorley, 2006) and PERMANOVA+ for PRIMER(Anderson et al., 2008). The data were fourth-root transformed todown-weight the dominance of abundant species, and then used togenerate a Bray–Curtis similarity matrix. Analysis of Similarity(ANOSIM) was performed to test for a priori differences betweeninvaded and non-invaded areas. In addition, semi-parametricpermutational analysis of variance (PERMANOVA) was used totest whether differences between invaded and non-invaded groupswere significant, using the similarity matrix created from fourth-root transformed data. A Type III sum of squares was employed withan unrestricted permutation of raw data.

Hierarchical clustering (using Bray–Curtis coefficients) and multi-dimensional scaling (MDS) were used to compare community structureamong sites. SIMPER (similarity percentage) analysis was performed(with a cut off of 90%) on transformed data to determine which specieswere most responsible for the dissimilarity between invaded and non-invaded groups. Taxa identified by SIMPER were tested for homogeneityof variances using Levene's test or by a graphical residual analysis (QuinnandKeough, 2002). In caseswhere varianceswere unequal, the dataweretransformed using either log, square-root or fourth-root transformations.Data were also tested for normality using the Kolmogorov–Smirnov test.In comparison of the biomass of species between invaded and non-invaded groups of sites, normality could not be achieved even aftertransformation of the data, so non-parametric Mann–Whitney U testswere used for these comparisons. Due to the large sample size, theMann–Whitney U statistic rapidly approached the normal distribution, soZ-values (normal distribution variate values) were reported along withp-values adjusted for ties.

2.5.2. CorrelationsAs an indication of the direction and strength of biological

interactions, Pearson Product–moment correlations were run betweenthe biomasses of functional groups. Abalone were excluded from theseanalyses because illegal fishing has diminished the population to suchan extent that their numbers are now unlikely to reflect biologicalinteractions.

2.5.3. Univariate analysesData on lobster abundance obtained from FIMS were not normally

distributed at the non-invaded stations and the variances betweeninvaded and non-invaded sites could not be equalized despite transfor-mation, because of numerous zero counts. A non-parametric Mann–Whitney U test was applied to test the significance of the differencebetween invaded and non-invaded stations, despite inequality ofvariance, so the analyses need to be interpreted in this light.

Biomasses of lobsters, scavengers, grazers, urchins, sessile taxa, kelps,understory algae and encrusting corallines were compared between sitesclassed as either ‘invaded’ or ‘non-invaded’ and among depth zones(b5 m, 6–12 m and 13–20 m). The data were tested for homogeneity ofvariances using Levene's test or a graphical residual analysis, and fornormality using a Kolmogorov–Smirnov test. In cases where varianceswere unequal, the data were transformed using log, square-root orfourth-root transformations. However, in the cases of scavengers, grazersand urchins, even after transformations the variances were still unequal,largely due to the abundance of zeros in the data. Despite the assumptionof equal variance being violated, statistical tests were still applied. Two-way factorial ANOVAs were run for Invasion Status×Depth and forDepth×Site. In cases of significant results, post-hoc Tukey tests wereapplied.

2.5.4. Species diversityTotal, floral and faunal species diversity were calculated for

each transect at each site for each depth using the ShannonDiversity Index H′=−∑ i=1

s pi(log epi), based on mean biomass asa measure of abundance. We then tested for a difference in meanspecies diversity between invaded and non-invaded sites at eachdepth zone. Data were first tested for homogeneity of variancesusing Levene's test and a graphical residual analysis; and fornormality using normal probability plots. Given that the assump-tion for normality (but not equality of variance) was violated, aMann–Whitney U test was applied.

Species dominance was assessed in two ways. The first employedSimpson's Dominance Index D=∑ i−1

s (pi)2. For the second, k-dominance plots were constructed by calculating the mean percentagebiomass for each species, ranking the species in order of importance, andplotting cumulative contribution to biomass against rank. From theseplots, a second index of dominance (D′)was derived:D′=1/S′, where S isthe inverse of the number of species contributing to 90% of the biomass.

3. Results

3.1. Spatial differences in the distribution of J. lalandii

Fisheries independent monitoring surveys (FIMS) revealed a signifi-cant difference in rock lobster CPUE (lobsters hoopnet−1soak period−1)between invaded and non-invaded stations (Fig. 3; adjusted Z=10.79,pb0.001). For this analysis, the data were averaged for all depths and allyears, but the same pattern was evident for all three depth rangesexamined and in each of the four years.

Size-frequency data for J. lalandii, recorded on inshore FIMS overthe years 2002–2005, showed that males (mean=78.8 mm CL,range=58–111 mm CL) were larger than females (mean=64.8 mmCL, range=44–94 mm CL) and that at least 50% of the overallpopulation exceeded 69 mm CL, the size at which J. lalandii canlegally be caught.

In addition to the FIMS, dive surveys showed that lobsters weresignificantly more abundant at invaded sites (see Fig. 5A, which isoutlined in more detail below). These two independent measures oflobster abundance confirmed our a priori classification of sites west andeast of Hermanus as being respectively ‘invaded’ and ‘non-invaded’.

3.2. Community composition

3.2.1. Multivariate analysesAn a priori ANOSIM based on the biomass of individual species

revealed significant differences in community composition betweeninvaded and non-invaded areas at each depth zone (b5 m: R=0.412,pb0.001; 6–12m: R=0.414, pb0.001 and 13–20 m: R=0.301,pb0.001). Semi-parametric PERMANOVA analyses also indicated signif-icant differences between Invasion Status (Pseudo-F=11.89, P(perm)b

0.001, SS=18,516), between Depth Zone (Pseudo-F=6.07, P(perm)b

0

2

4

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61 62 63 64 65 68 69 70 71 77 78 79 80 84 85 86 87 88 90Mea

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At 21-30 m

At 11-20 m

At 0-10 m

FIMS CPUE trap data

***

Fig. 3. Mean (+ 1 SE) CPUE for Jasus lalandii (lobsters hoopnet−1 soak period−1) for Fisheries Independent Monitoring Surveys (FIMS) in the area EOCH, averaged across all depths forthe period 2002–2005. Circles showmean values for each depth zone. The six underlined stations are where benthic samples were taken: 61= Hangklip, 63= Betty's Bay, 69=MudgePoint, 78 = Romans Bay, 79 = Kruismans Bay, 90 = Quoin Point. ***=pb0.001 for a Mann–Whitney U-test between invaded and non-invaded areas.

37L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

0.001, SS=18,907) and an interaction between Invasion Status×DepthZone (Pseudo-F=2.54, P(perm)b0.001, SS=7915). Pairwise tests indicat-ed that each of the depth zones was significantly different (b5 m vs6–12 m: t=1.94; b5 m vs 13–20 m: t=3.17; 6–12 m vs 13–20 m:t=2.09; P(perm)b0.001 in all cases). MDS analysis showed that in allthree depth zones, the invaded and non-invaded sites separated out,with the exception of Kruismans Bay which tended to cluster withthe invaded sites (Fig. 4). This outcome was also evident in a clusteranalysis (not shown).

3.2.2. Characteristic speciesAt all three depth ranges, dissimilarity between invaded and non-

invaded sites was high, spanning 75.8–86.4% and averaging 76.9%.Species characterizing the two states at all three depths yielded severalclear patterns, as evidenced for all three depth zones (Appendix A: Figs.Ai–Aiii). First, predators and macroalgae were significantly moreabundant at invaded sites. Second, herbivores and encrusting corallineswere less abundant at invaded sites, most obviously so at the twoshallower depth ranges in the case of the corallines (Appendix A: Figs.

3D Stress: 0.13

R

R

R

R

R

R

Q

Q

Q

Q

Q

QQ

QQ

R

RR

K

KK K K

K

K K

K

B

M

M

M

BB

MM

M

Fig. 4. MDS plot based on biomass data for species, averaged for each transect at each of thseparates samples distinguished at 80% dissimilarity.

Ai, Aii). Strikingly, the urchin P. angulosuswas consistently absent fromall invaded sites, but present at all non-invaded sites — where itsometimes achieved a considerable biomass. The abalone H. midaewastoo scarce to contribute to the distinguishing taxa. Adults were rare atboth invaded and non-invaded sites (averaging respectively 7.3 and7.5 g m−2), and juveniles were completely absent at invaded sites andrecorded in only one out of a total of 162 quadrats at the non-invadedsites. Third, sessile animals showed mixed responses, but overall weremore abundant at invaded sites. Fourth, depth-related patternsemerged. Differences between invaded and non-invaded sites notedabove were consistent across depths for all functional groups exceptsessile animals, which were more abundant at invaded than non-invaded sites in themost shallow depth range (Appendix A: Fig. Ai), butless abundant there in the two deeper zones (Appendix A: Figs. Aii, Aiii).The average number of species contributing to 90% of the dissimilarity,integrated across all three depth intervals, was 18, but rose from 11 atb5 m through 14 at 6–12 m to 26 at 13–20 m, with the most noticeablecontributors to this increase being macroalgae and sessile animals(Appendix A: Figs. Ai–Aiii).

<5m6-12m13-20m

<5m6-12m13-20m

Invaded

Non-invaded

B M

MM

H

H

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H = Cape HangklipB = Betty’s BayM = Mudge PointR = Romans BayK = Kruismans BayQ = Quoin Point

ree depth intervals, for three ‘invaded’ and three ‘non-invaded’ sites. The dashed lined

38 L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

Overall, invaded sites had a greater biomass of predators, macro-algae and sessile species relative to non-invaded sites; whereas non-invaded sites had more grazers and encrusting corallines relative toinvaded sites.

3.2.3. Univariate analyses of functional groupsLobster biomass was significantly greater at invaded sites (Fig. 5A),

while at non-invaded sites lobsters were either rare or absent. There

Invaded Non-invaded

<5 m6-12 m13-20 m

0 = absence

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Fig. 5. Mean biomasses (+ 1 SE) of functional groups at ‘invaded’ and ‘non-invaded’ sites in(right). CH = Cape Hangklip, BB = Betty's Bay, MP = Mudge Point, RB = Roman's Bay, KB

was no consistent depth pattern at invaded sites, but at non-invadedsites, lobster biomass was always greatest in the deepest depth-zone(13–20 m), although only significantly so at Romans Bay (Fig. 5A).

In contrast to lobsters, scavengers, grazers and urchins were allsignificantly more abundant at non-invaded sites (Fig. 5B–D), withscavengers and grazers rare at invaded sites (Fig. 5B, C), and urchinscompletely absent (Fig. 5D). While scavengers showed no consistentdepth pattern (Fig. 5B), both grazers and urchins declined with depth

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2000

2500

3000

Bio

mas

s (g

.m-2

)

*** a

b b

a

a

b

0

20

40

60

80

100

120

140

160

B) Scavengers

Bio

mas

s (g

.m-2

)

***

aa a a a

aa a

a

a

b ba a

a a

ab

b

*** ***

*** ***

*** *** *** *** *** *

*** ** ***

three depth zones. Dashed lines separate ‘invaded’ sites (left) from ‘non-invaded’ sites= Kruismans Bay and QP = Quoin Point.

39L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

at non-invaded sites, except for Kruismans Bay where grazers wereless abundant than at other non-invaded sites (Fig. 5C), and urchinswere unexpectedly absent in the shallow depth zone (Fig. 5D).

Sessile animals (Fig. 5E) were relatively uniform in their abundance.Therewasno significant difference in their biomass between invaded andnon-invaded areas and no consistent depth pattern. In the b5-m depthzone at Mudge Point, their biomass was extremely high (20,340 g m−2),largely accounting for the greater pooled biomass of sessile species in theshallows of invaded sites.

Kelps (Fig. 5F) were significantly more abundant at invaded sitesand their biomass decreased with depth and were always lowest in thedeepest depth zone. Understory algae (Fig. 5G) were also significantlymore abundant at invaded sites, but surprisingly showed no patternwith respect to depth.

Encrusting corallines (Fig. 5H) had a significantly smaller biomass atthe invaded sites and showed no depth pattern there. At the non-invaded sites, they decreased with depth, with very low values in the13–20 m depth zone, where there was no distinction between invadedand non-invaded sites. For all three algal groups, Kruismans Bay againdiffered from the other non-invaded sites, being intermediate betweenthe condition at invaded and other non-invaded sites.

3.3. Species diversity

At all three depths, faunal diversity was greatest in the non-invaded sites (Fig. 6A), whereas floral species diversity was greatestin the invaded sites (Fig. 6B), but these trends were non-significant inhalf the cases. Both faunal and floral species diversity increased withdepth, although irregularly so in the case of fauna at non-invadedsites.

Total, faunal and floral species dominance (as reflected by theSimpson Dominance Index D and the inverse of the number of species

Invaded

Non-invaded

*** ** NS

NS**

NS

Depth Zones

Depth Zones

B) Flora

A) Fauna

Sha

nnon

Div

ersi

ty In

dex

(H’)

Sha

nnon

Div

ersi

ty In

dex

(H’)

Fig. 6.Mean (+ 1 SE) values of (A) faunal and (B) floral species diversity (based on theShannon Diversity Index, H′ for invaded and non-invaded areas in three depth zones).Mann–Whitney U-tests: **=pb0.01 and ***=pb0.001; NS = not significant.

accounting for 90% of the biomass, D′), decreased with depth at bothinvaded and non-invaded sites (Table 1). Although this pattern wassometimes obscure between the shallow and intermediate depths, itwas particularly obvious when these were compared to the 13–20-mdepth, which had the lowest values in all but one case (Table 1).Faunal species dominance was always greater at invaded sites,whereas floral species dominance was greater at non-invaded sitesin all instances bar one (Table 1). These trends were the converse ofthose associated with measures of diversity.

3.4. Biotic correlations

Pearson Product–moment correlations were used to examine thestrength and direction of relationships between functional groups,summarized in the form of impact graphs. In the b5 m depth zone(which we used for illustrative purposes; Fig. 7), the rock lobster J.lalandii had significant negative impacts on the urchin P. angulosus,grazers and encrusting corallines. It also had non-significant positiveimpacts on macroalgae (kelp, foliar and turf) and an unexpectedsignificant positive association with sessile species. Scavengers werestrongly positively associated with urchins, but probably for theindirect reason that both were consumed by rock lobsters, bothbeing common only where lobsters were rare. Urchins had asignificantly positive relationship with encrusting corallines and aprobably indirect, but also significant, positive association withgrazers. Urchins were additionally negatively associated with kelp,foliar algae and turf, albeit significantly so only in the case of kelp.Grazers were non-significantly negatively correlated with allmacroalgal groups and sessile species, and also appeared to promoteencrusting corallines. The three macroalgal groups (kelps, foliar andturf) were negatively associated with urchins, grazers and encrust-ing corallines, and kelp specifically was significantly associated withsessile species. Encrusting corallines were positively associated withurchins and grazers, and negatively associated with all three groupsof algae — although only significantly so in the case of kelp. Sessilespecies had a significant negative relationship with encrustingcorallines.

Data for the two deeper depth zones (Appendix B: Figs. Bi–Bii),tended to show the same patterns, with three departures. First,compared to the 12 significant correlations that existed at b5 m,there were only nine at 6–12 m, and only five at 13–20 m.Most of thediminishment in significant correlations was due to urchins andkelps becoming too scarce in deeper water for meaningful relation-ships to exist, but there was also a substantial decline in the numberof significant associations among other groups. Second, while J.lalandii was consistently negatively correlated with grazers andscavengers and positively correlated with foliar and turf algae at alldepths, at 13–20 m it became positively associated with encrustingcorallines and negatively associated with sessile species. Third, foliaralgae had a significant negative correlation with sessile species in thedeepest zone. This was intuitively more logical than the positive

Table 1Dominance, measured as D (Simpson's Dominance Index) and as D′ (the inverse of thenumber of species accounting for 90% of the biomass), expressed for total, faunal orfloral components in three different depth zones at invaded and non-invaded sites.

Invasionstatus

b5 m depth 6–12 mdepth

13–20 mdepth

D D′ D D′ D D′

Total Invaded 0.44 0.25 0.29 0.10 0.21 0.05Non-invaded 0.39 0.20 0.41 0.17 0.26 0.06

Faunal Invaded 0.79 0.66 0.52 0.40 0.58 0.25Non-invaded 0.41 0.17 0.41 0.20 0.41 0.13

Floral Invaded 0.63 0.33 0.46 0.13 0.26 0.08Non-invaded 0.56 0.50 0.61 0.33 0.33 0.13

Rock lobster

Impacted group

Imp

acti

ng

gro

up

0

+ 1.00

Enc

rust

ing

**

Fol

iar

Tur

f

Kel

p

Gra

zer

*

Sca

veng

er

Urc

hin

*

Ses

sile

s

*

Urchin**

*******

NA NA

Grazer**

NA

* ***

Encrusting NA****

NA

**NA

Sessiles **NANANA NA

FoliarNANA

Kelp **NANA ***

Turf NANA

Scavenger NANA NANANA

*** *

-1.00

Fig. 7. Relationships between selected groups at a depth of b5 m. Bars indicate the correlations between groups (scaled between 0 and ±1 as per the top left scale). Positivecorrelations are shown in black above the zero line for each group, and negative correlations in gray below. Impacted groups are listed on the horizontal axis and groups responsiblefor the impact on the vertical axis. Significant correlations are indicated as follows: *=pb0.05, **=pb0.01 and ***=pb0.001; NA = not applicable.

0.0

0.5

1.0

1.5

2.0

0.0 0.1 0.2 0.3 0.4 0.5

Rock lobsters (log number.m-2 + 1)

Urc

hins

(lo

g nu

mbe

r.m

-2 +

1)

a

Conditions

Sta

te

Conditions

Sta

te

Conditions

Sta

te

(A) (B) (C)

b

c

Fig. 8. Relationship between the density of urchins and rock lobsters (log numbersm−2+1), showing adiscontinuous shift between invaded andnon-invaded sites. Regressionlines were fitted to data from non-invaded sites (black diamonds) and invaded sites (graysquares). Kruismans Bay was grouped with the invaded sites due to the scarcity of urchinsthere, and a benthic community that was more similar to invaded sites. Insets illustrate thethree different ways in which state shifts can take place in an ecosystem in response tochanges in conditions: (A) smooth, (B) abrupt or (C) a discontinuous regime shift (Redrawnfrom Scheffer and Carpenter, 2003). Point a indicates the rock lobster density (0.22 rocklobsters m−2) abovewhich therewere no urchins, and b indicates the rock-lobster density of0.05: between a and b urchins were present but scarce; c indicates high densities of urchins(>20m−2) in the absence of lobsters.

40 L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

nature of the associations between sessile species and both J. lalandiiand foliar algae in shallower zones.

Because of the ecological importance of lobster–urchin interactions,the relationship between their densities was explored more fully, andwhen regressed against each other, showed a discontinuous shiftbetween sites with high densities of urchins and those that lacked orhad low densities of urchins (Fig. 8). Above 0.22 lobsters m−2, urchinswere consistently absent, and at 0.05–0.22 lobsters m−2, urchinstended to be present but scarce. Only in the absence of lobsters didthey attain substantial densities in excess of 20 urchins m−2.

4. Discussion

Our study addressed three primary questions. (1) Was there asignificant difference in J. lalandii density between putative invaded andnon-invaded areas? (2) Did benthic community composition differbetween invaded and non-invaded areas in different depth zones? (3)What were the likely causes of emergent patterns?

Using these questions as a focus,we first summarize the outcomes ofour research (Fig. 9) and compare them with our a priori expectations(Fig. 1), and then broaden their application to the larger issues of top-down effects of predators, overfishing, climate change and alternativestable states.

4.1. Spatial differences in J. lalandii abundance

We hypothesized that J. lalandii would be significantly moreabundant at invaded sites than non-invaded sites. Fisheries Indepen-dent Monitoring Surveys (FIMS) and diver surveys independentlyconfirmed our a priori classification of sites as either ‘invaded’ or ‘non-

invaded’. At least half the population was ≥69mm CL, coinciding withboth the 70-mm legal minimum size for commercial capture and theminimum size J. lalandii must achieve to consume the urchin P.angulosus (Mayfield et al., 2001).

Invaded state Non-invaded state

Jasus lalandii

Grazers

Macroalgae

Encrustingcoralline

Sessile animalsScavengers

Jasus lalandii

Macroalgae

Encrustingcoralline

Sessile animals

Scavengers

Urchins

Grazers

Kelp

Foliar

Turf

Kelp

Foliar

Turf

110.5 5.4

24.6 260

899

118195032

1283 5382

4.8

22.9

7708.941062

+

-Predation -

Predation

+Unknowncause

+Indirect-

Predation

-Competition

for space -Shading

-Competition

for space

-Grazing

+Indirect

-Grazing

-Competition

for space-Competitionfor space

-Disturbance

+

Removal of competitive algae

+Settlement on substratum?

Urchins0.0

Fig. 9. Interaction webs for invaded and non-invaded states in the b5 m depth zone. The bold arrows imply strong effects and the non-bold arrows weaker effects. Causes of positive(+) and negative (−) effects are indicated in boxes. Mean wet biomasses (g m−2) are indicated by numbers and the areas of the circles, which are proportional to biomass on a logscale.

41L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

4.2. Were benthic communities different between invaded and non-invaded areas?

Given the reputation of rock lobsters for altering communitystructure (Babcock et al., 1999; Barkai and Branch, 1988b; Robles etal., 1990), we expected that in areas where J. lalandii had invaded andachieved high densities, the community would differ from that inareas where lobsters have remained rare.

Indeed, multivariate analyses revealed significant differences in allthree depth zones. The relative abundances of functional groups withininvaded and non-invaded kelp forests and their deduced interactions(summarized in Fig. 9) distil to the following patterns: (1) Conformingto expectations, J. lalandiiwas 20 timesmore abundant at invaded areas,where it was strongly negatively associated with urchins, grazers andencrusting corallines and, to a lesser extent, scavengers. (2) Herbivores(grazers and urchins) were reduced at invaded sites. From our a prioriflow model (Fig. 1) we expected urchins to be scarce at invaded sites,but the situation was more extreme and they were absent altogether,whereas they dominated the herbivore component at non-invadedsites. Grazers were also an order of magnitude more abundant at thenon-invaded sites. (3) Macroalgal biomass at invaded sites was morethan double that of non-invaded sites. (4) In contrast, encrustingcorallines were more abundant at non-invaded than invaded sites. (5)Sessile species appeared to be more abundant at invaded sites.However, this was due to a single very high value and they wereotherwise relatively uniform across sites and depths.

The non-invaded Kruismans Bay site had a greater abundance of J.lalandii, kelp and understory algae relative to the other two non-invaded sites, and P. angulosus was unexpectedly rare or absent. Thismay explain its anomalous grouping with the invaded Betty's Bay site.The relationship between urchin densities and those of rock lobstersshowed a discontinuous shift between invaded and non-invaded sites(with the exception of Kruismans Bay for reasons outlined above). For awide range of lobster densities above a threshold of 0.22 lobsters m−2,urchins were absent; but below that threshold, urchins were present at

widely varying densities that are likely to have been influenced byadditional factors such as recruitment, wave action etc. Similar patternshave been observed between urchins and brown algae along the westcoast of Vancouver Island (Watson and Estes, 2011). Mayfield andBranch (2000) reported that at a critical density ≥0.25 lobsters m−2

very few, if any, urchins would survive. Our threshold value of 0.22coincides closely with this. At Kruismans Bay, lobsters reached a meandensity that may have been sufficient to explain the absence or scarcityof urchins there. The intermediate position of Kruismans Bay betweeninvaded and non-invaded sites therefore most likely reflects the factthat rock-lobster abundances there were approaching levels at whichthey strongly influenced other elements of community composition.

4.3. Depth-related patterns

At both invaded and non-invaded sites, benthic communities (andspecies interactions) changed with depth. As depth increased, mostdifferences between invaded and non-invaded areas were retained, butsomedepth-related differenceswere superimposed. J. lalandii increasedwith depth. Urchins declined with depth at sites where they werepresent. Kelps diminishedwith depth,while understory algae increasedin diversity although their biomass surprisingly fluctuated unpredic-tably with depth. The prevalence of encrusting corallines at non-invaded relative to invaded sites disappeared at 13–20 m as theybecame less abundant with depth. The fact that they declined withdepth is contrary to the predictions of Steneck and Dethier (1994), whoargued that as productivity diminishes with a reduction in light,encrusting corallines will become proportionally more dominant.

There were fewer significant interactions between trophic groups asdepth increased (12, 9 and 5 at respectively 0–5, 6–12 and 13–20 m),but therewere progressivelymore taxa contributing to the dissimilaritybetween invaded and non-invaded sites (11, 14 and 26 in theserespective depths). The number of species with significantly differentbiomasses between invaded and non-invaded sites similarly increasedwith depth, being respectively 7, 10 and 19 (Appendix A: Figs. Ai–Aiii).

42 L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

Species dominance also declinedwith depth (Table 1). Thus, an increasein depth led to more heterogeneous benthic communities and lessdominance, and to a narrowing of the divide between invaded and non-invaded sites. Nevertheless, it seemed that the presence or absence ofhigh densities of lobsters set the primary ecosystem state and thatdepth-related differences among communities were then superim-posed onto this lobster effect.

4.4. Explaining patterns in community structure

Biotic interactions that potentially explain the observed patternsin community structure across invasion status and depth include (1)direct predation, indirectly altering (2) herbivory, (3) competitionand (4) facilitation.

4.4.1. Rock lobsters: a notorious predatorDepending on their density and size structure, rock-lobster popula-

tions have the ability to radically alter community structure andfunctioning. This has been reported both elsewhere in the world(Babcock et al., 1999; Robles, 1987; Robles et al., 1990; Tegner andDayton, 1981, 1991; Tegner and Levin, 1983) and at other localities inSouth Africa (Barkai and Branch, 1988a,b; Barkai and McQuaid, 1988),and is underpinned by rock lobsters being opportunistic foragers thatfeed on awide variety of prey (Barkai and Branch, 1988a,b; Barkai et al.,1996; Cox et al., 1997; Edgar, 1990; Goñi et al., 2001; Mayfield andBranch, 2000; Mayfield et al., 2000a,b). In the absence of preferred fooditems, J. lalandii is capable of switching to unusual items (e.g. barnaclerecruits), and is thus able to maintain high densities even in areas ofapparently low food abundance (Barkai and Branch, 1988c). On thewest coast of South Africa, J. lalandii has large negative effects on itsmain prey, leading to the development of alternative stable states ofcommunity composition in localities where it is abundant compared toareas where it is rare (Barkai and Branch, 1988a,b; Barkai andMcQuaid,1988). The invasion of J. lalandii into the area EOCH coincided with anabrupt decline in preferred prey items such as P. angulosus and otherherbivores (Blamey et al., 2010; Mayfield and Branch, 2000; Tarr et al.,1996).

4.4.2. Herbivore–algal interactionsThe absence of P. angulosus and the low biomass of other herbivores

in invaded areas correspondedwith a high biomass of understory algaeand kelps. The kelp biomass in invaded areaswasmore thandouble thatof the non-invaded areas in the b5-m depth zone; and the same wastrue for understory algal biomass in all depth zones.

At sites where rock lobsters were scarce or absent, P. angulosus wasthe dominant subtidal herbivore and there was also a markedly greaterbiomass of other herbivores such as the gastropods T. sarmaticus, T.cidaris and Oxystele sinensis. Urchins are renowned worldwide for theirgrazing abilities (Andrew, 1993; Andrew and Underwood, 1993;Carpenter, 1986, 1988; Hay, 1984; Lawrence, 1975; Mann, 1977) andtheir capacity to transform macroalgal-dominated systems into coral-line barrens (Breen and Mann, 1976; Dayton et al., 1998; Estes andDuggins, 1995; Mann, 1977; Tegner and Dayton, 1991; reviewed inSteneck et al., 2002). However, it seems unlikely that P. angulosus alonecan achieve this. Firstly, P. angulosus feeds primarily on drift kelp andonly in the absence of drift does it become an active grazer (Fricke,1979; Velimirov and Griffiths, 1979; Velimirov et al., 1977). Day andBranch (2002a) showed that experimental removal of P. angulosus onthe Cape Peninsula near where our work was done did not lead to anysignificant changes in the abundance of foliar algae, kelp sporelings orencrusting corallines, and they attributed this to the fact that P.angulosus traps drift algae rather than acting as a grazer. Secondly,turbulent waters in the coastal zone prevent P. angulosus fromascending kelp plants to graze on them. Anderson et al. (1997) have,however, suggested that should events such as large-scale storms orunsustainable harvesting lead to depletion of kelp beds, then high

densities of P. angulosus and other grazers in combination mightprevent re-colonization by kelp, resulting in coralline-dominatedbarrens. Our study showed significant positive correlations betweenP. angulosus and encrusting corallines, and negative correlationsbetween P. angulosus and macroalgae. Moreover, the fact that algalabundance in non-invaded areas was less than half that in invadedareas, where urchins were absent and the biomass of grazers waslow, suggests that high densities of urchins and other herbivores actin combination to control kelp and understory algal abundance.

Although our data are correlative, they do constitute strongcircumstantial evidence that high densities of rock lobsters depletethe total pool of herbivores, indirectly leading to proliferation ofmacroalgae, in turn diminishing encrusting corallines because theybecome overgrown by macroalgae. This may have repercussions forH. midae, as its recruits settle specifically on encrusting corallines(Day and Branch, 2000a). The correlative patterns we describe, andour attribution of them to rock lobster abundance, gain credencefrom the fact that corresponding changes in community compositionemerged from temporal comparisons before and after the ‘invasion’by lobsters (Blamey et al., 2010).

4.4.3. Algal–algal interactionsAlthough kelps achieved a high biomass at invaded sites and spanned

a wide range of biomass across all sites and depths, there was nosignificant negative correlation between them and understory algae, sokelps did not appear to influence understory algae in the manner seenelsewhere (e.g. Clark et al., 2004; Dayton et al., 1984, 1992; Edwards,1998; Kennelly, 1989; Melville and Connell, 2001; Reed and Foster,1984; Watson and Estes, 2011). However, there was a significantnegative correlation between kelp andencrusting corallines. In Australia,experiments have shown that a canopy cover of Ecklonia radiata reducesunderstory (foliar/turf) algae but increases the abundance of encrustingcorallines (Melville and Connell, 2001). A similar association has beenreported in California (Reed and Foster, 1984). In our study, differentpatterns and processes emerged — there was a decrease in encrustingcorallines at invaded areas and an increase in understory algae. It seemslikely that intense predation by rock lobsters in invaded areas directlyreduces herbivores there, indirectly leading to a cascading positive effecton macroalgae and thence to a negative effect on encrusting corallines.Predation by lobsters on encrusting corallines (presumably as a source ofcalcium) has previously been recorded (Mayfield et al. 2000a,b), andmayhave contributed to their depletion in invaded areas. Further studiesare needed to determine what role canopy-cover plays, and whether itinfluences the species composition of understory algae, even if it doesnot appear to affect their overall biomass.

Encrusting corallines are both resistant to, and tolerant of, urchingrazing (Breitburg, 1984) and inmany cases they indirectly benefit fromgrazing through the removal of erect algae that would otherwisesmother them (Andrew and Underwood, 1993; Breitburg, 1984; Bulleriet al., 2002; Fletcher, 1987; Steneck, 1982). In addition, experimentalwork has shown that following the removal of erect algae, encrustingcorallines can exert a negative effect on their recruitment, out-competingthem for space and thus retarding macroalgal re-colonization (Bulleri etal., 2002; Johnson and Mann, 1986; Keats et al., 1994, 1997). Werecorded a strong positive correlation between encrusting corallines andboth urchins and grazers, implying that encrusting corallines indirectlybenefit from herbivores. This is not surprising, given that numerousstudies have shown that in productive environments where herbivoresare abundant, encrusting corallines dominate (Fletcher, 1987; Lawrence,1975; Menge and Lubchenco, 1981; Paine, 1984; Steneck, 1986; Steneckand Dethier, 1994). Furthermore, experimental work has shown thaturchins and other grazers are often required tomaintain subtidal beds ofencrusting corallines (Fletcher, 1987; Littler et al., 1995;Morrison, 1988).In our study, the relationships among different functional groups of algaewere, however, not straightforward or consistent. Encrusting corallineswere (non-significantly) negatively correlated with both turf and foliar

43L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

algae in the two shallowest zones but positively correlated with them inthe deepest zone. Their relationship with kelp was more clear-cut, beingconsistently negative. This was probably underpinned by the negativeeffects of herbivores on kelp, but we cannot rule out the possibility thatencrusting corallines themselves exert a direct negative effect on erectalgae.

4.4.4. Urchin–abalone interactionsIn invaded areas EOCH, the urchin P. angulosus was entirely absent,

whereas it was abundant both in non-invaded areas and, prior to theinvasion, in areas fated to become invaded (Blamey et al., 2010). Itsdecline thus coincides spatially and temporally with the arrival of highdensities of J. lalandii. The depletion of P. angulosus has seriousconsequences for the abalone H. midae, which has declined drasticallysince themid 1990s due to both overfishing caused by poaching (Hauck,2009; Hauck and Sweijd, 1999), and its critical dependence on urchins(Day and Branch, 2000a,b). Associated with the decline of urchins,abalone recruitment in this area dropped to exceptionally low levels(Tarr et al., 1996). Nevertheless, we anticipated finding juvenile abalonesheltering under theurchins at non-invaded siteswhere urchin densitieshave remained high. Prior to the invasion, Tarr et al. (1996) recordeddensities of juveniles of about 40–60 m−2 at all sites, whereas after theinvasion the densities dropped to zero at sites that had been invaded,while remaining at about 40–60 m−2 at non-invaded sites. Surprisingly,we recorded virtually no juvenile abalone at any of the sites, irrespectiveof whether they were invaded or non-invaded. This strongly suggeststhat recruitment failure is taking place: that adult stocks have beendepleted to the point that reproduction and subsequent recruitment arecollapsing. As broadcast spawners, abalone will be prone to the Alleeeffect (Gascoigne and Lipcius, 2004), whereby populations need toachieve minimal densities for fertilization to be effective in the face ofdilution of sperm and eggs in the water column. Population collapse ofbroadcast-spawning urchins has previously been attributed to this effect(Quinn et al., 1993).

4.5. Species diversity

Floral species diversity was, on average, greater at invaded siteswhereas faunal diversity was greater at non-invaded sites. Both floraland faunal diversity increasedwith depth. Species dominance displayedthe opposite trends to diversity.

These trends have several possible explanations. Firstly, top-downpredation may contribute to differences in faunal diversity. Predationcan increase diversity (Menge and Sutherland, 1976; Paine, 1966) if itprevents monopolization by competitively dominant species. How-ever, in the invaded system, where rock lobsters were abundant,faunal diversity was diminished, probably because intense predationhas left only a subset of species that are disfavored as prey and haveassumed greater dominance. Dietary studies confirm that in invadedareas, J. lalandii consumes smaller and less diverse prey with a lowerenergy content than in non-invaded areas where its densities are low(Haley et al., 2011).

Secondly, depth reduces light and productivity, and we expectedalgal diversity to decrease with depth. However, this was not the case:floral diversity was lowest in the shallows and increased with depth inboth invaded and non-invaded areas. Similar depth-related patternshave been observed in coral communities (Glynn, 1976; Huston, 1985a,b;Loya, 1972). Three different causes may explain the observed trends:

(1) It is likely that shallow depths associated with high productivity(Steneck and Dethier, 1994) are monopolized by a few species and thatwith an increase in depth, dominance by any one species is reduced,allowing species diversity to increase. Indeed, for floral species,dominancewas greatest in the shallows,where a fewspecies contributedthe bulk of the biomass, and diminished with depth. Floral speciesdominancewas also greater at non-invaded sites, possibly because a fewspecies are resistant to grazing and have thus established dominance.

(2) Wave action decreases with depth, with the most intense physicaldisturbance occurring in the shallows. The intermediate disturbancehypothesis (Connell, 1978; Grime, 1973; Sousa, 1979) proposes thatmaximum diversity will occur at intermediate levels of disturbance.However, disturbance is often difficult to quantify (Reynolds et al., 1993)and the term “intermediate levels” is rather subjective, often beingdefined circularly in terms of the conditions that produce maximumdiversity (Huston, 1994; Shea et al., 2004). One might expect speciesdiversity to be enhanced in the shallow depths, where waves generatephysical disturbance. However, if the disturbance there is frequent andsevere, wave action may favor only those species that tolerate suchconditions. (3) Floral diversity in the invaded and non-invaded areas wasrelatively similar in the shallows. However, as depth increased, it becamegreater in the invaded areas. This could reflect a prevalence of physicalcontrol in the shallows, giving way to greater biological controls as waveaction subsides with depth and dominance diminishes. Herbivorebiomass was significantly lower in the invaded areas than in the non-invaded areas, and although increased grazingpressure can increasefloralspecies diversity (Hily et al., 1992), it is also known to decrease diversity(Duggins and Dethier, 1985; Himmelman et al., 1983; Vance, 1979;Wootton, 1995). This could explain why non-invaded areas, whereherbivores were significantly more abundant, supported a lower floraldiversity.

All of these potential effects of depth, predation and grazing (andtheir interactions) are hypothetical and based on observational compar-isons, but are amenable to experimental tests and modeling.

4.6. Top-down control, overfishing and climate change

Many marine ecosystems are viewed as resource-controlled(bottom-up control). For instance, nutrient-limitation of phyto-plankton dynamics can powerfully influence pelagic systems (e.g.Frederiksen et al., 2006; Richardson and Schoeman, 2004; Ware andThompson, 2005). The opposite of this – systems that are consumerdriven (top-down control) – has often been thought to be limited tonearshore and intertidal ecosystems (Frank et al., 2007). Top-downcontrol has frequently been documented in coastal kelp-forestecosystems (Estes et al., 1998; Halpern et al., 2006; Steneck et al.,2002; Watson and Estes, 2011), subtidal temperate reefs (Barrett etal., 2009; Götz et al., 2009a,b; Guidetti, 2006; Shears and Babcock,2002, 2003), and intertidal rocky shores (Paine, 1974, 1980). In thepast decade, evidence for top-down control in exploited ecosystemshas risen to prominence as overfishing of top predators has reachedcritical levels (e.g. Frank et al., 2005; Worm and Myers, 2003).Strong (1992) suggests that top-down structuring of ecosystems isnot typical, but instead indicates a form of biological instability, asreflected in over-fishing that has led to the transformation of manycoastal ecosystems (Jackson et al., 2001). For example, the removal oftop-predators through over-fishing is reported to disrupt predator–prey relationships, allowing prey to then consume the early stages oftop-predators or to compete with them for food, thereby inhibiting therecovery of the predators (Köster and Möllmann, 2000; Swain andSinclair, 2000). A combination of both top-down and bottom-up controlis known as wasp-waist control (Rice, 1995). This occurs when speciesat an intermediate trophic level influence other species at both higherand lower trophic levels — a classic example being small pelagic fish,which exert a top-down control on zooplankton as well as havingbottom-up influences on predators (Cury et al., 2000).

Along the south-western Cape coast where our work was done,overfishing of predatory linefish has led to their collapse (Griffiths,2000). Fishing success and the size of fish captured have declined inboth commercial and recreational fisheries, and catches have shiftedfrom high-value, slower-growing species to low-value, short-livedspecies (Attwood and Farquhar, 1999). It is possible that the depletionof these predatory fish created an unstable ecosystem and opened anopportunity for range expansion by J. lalandii. Alternatively, or in

44 L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

combination, changes in physical conditions could have allowed J.lalandii to increase in density EOCH.

In south-east Australia, increased poleward penetration of theEast Australian Current has resulted in a 1.5 °C increase in sea surfacetemperature along eastern Tasmania (Ridgway, 2007). This warmingallowed the long-spined sea urchin Centrostephanus rodgersii toextend its range southwards from New South Wales to easternTasmania (Ling et al., 2008, 2009a). Consequences of this range-extension are potentially catastrophic for the following reasons. (1)C. rodgersii over-grazes macroalgal habitats, creating and maintain-ing alternative stable ‘barren’ states (Andrew, 1993; Andrew andUnderwood, 1993; Fletcher, 1987). (2) Barrens formed by C. rodgersiialong Tasmania have resulted in a significant decline in macroalgalcover and habitat-forming invertebrates and hence a loss of speciesabundance, richness and diversity (Ling, 2008). (3) Given thewarming of coastal waters, C. rodgersii is currently capable ofreproductively maintaining itself along eastern Tasmania (Ling etal., 2008). (4) The rock lobster Jasus edwardsii – amajor predator of C.rodgersii – has been over-fished in Tasmanian waters, resulting inboth fewer and smaller-sized lobsters, diminishing consumption ofC. rodgersii and heightening the impact of this urchin on reef habitats(Ling et al., 2009b).

Rouault et al. (2009) have reported a significant increase in the seasurface temperature of the Agulhas Current system since the 1980s,associated with an increase in wind-stress curl in the South IndianOcean, which is consistent with a poleward shift in westerly windsreported elsewhere (Gillett et al., 2003; Seidel et al., 2008; Thompsonand Solomon, 2002). Simultaneous with the warming of offshorewaters, inshore coastal waters east of Cape Agulhas have cooled alongthe south coast since the 1980s (Rouault et al., 2009; Roy et al., 2007).This cooling may have contributed to observed eastward shifts inkelps (personal observations), rock lobsters (Blamey et al., 2010;Cockcroft et al., 2008; Mayfield and Branch, 2000; Tarr et al., 1996),and pelagic fish (Fairweather et al., 2006), all of which are cold-waterspecies; and a range retraction of the warm-water brown musselPerna perna (Mead, 2011). However, the fact that inshore coolingoccurred predominantly east of Cape Agulhas (ca. 100 km east ofwhere we worked) argues against it being the cause of the lobsterinvasion EOCH, and attempts to link geographic shifts of species inthe region to changes in environmental conditions have beenproblematic because of inadequate long-term datasets (Cockcroft etal., 2008).

As Ling et al. (2009b) have described, range-expansion induced byclimate change, coupled with ecosystem instability through the over-fishing of top-predators, does not bode well for the maintenance ofecosystems or their ability to recover from perturbations. Coupling (1)the eastward shift in J. lalandii into an unstable over-fished systemwith(2) the notorious ability of this lobster to alter benthic communities(Barkai and Branch, 1988a,b; Barkai and McQuaid, 1988), it is highlylikely that J. lalandii has initiated a regime shift in the communitystructure at invaded sites EOCH, through top-down predatory control.

4.7. A case for alternative stable states?

Regime shifts can be categorized as smooth, abrupt or discontinuous(see Fig. 9 inset), the last being synonymous with an alternate stablestate. Several authors have specified criteria to evaluate whether achange in an ecosystem can justifiably be called a ‘regime shift’ (Collie etal., 2004; deYoung et al., 2004; Scheffer and Carpenter, 2003). Thesecriteria require that a shift be low frequency, high amplitude, abrupt,prolonged, regional, and affect more than two trophic levels; and that itmust not be maintained by human influence. Lees et al. (2006) havegone on to quantify some of these criteria on the basis of previous casesregarded as regime shifts, which collectively had aminimum amplitudeof 46%, occurredwithin 1–7 years, persisted for aminimumof ten years,and affected at least three trophic levels. They also highlight that such

shifts can be driven not only by climate forcing, but also by biological oranthropogenic forcing, e.g. range expansion or overfishing.

Ling et al. (2009b) have provided arguably one of the best cases ofa regime shift, demonstrating a link between physical forcing (seasurface temperature rise) and a biological response (range expansionof the urchin C. rodgersii), and showing that the ecosystem effects ofthis response span 2–3 trophic levels. In addition, they showed howfishing effects compound the situation (absence of large lobsterspromoting persistence of urchins).

Along the South African west coast, Barkai and Branch (1988a,b) andBarkai and McQuaid (1988) described a localized alternate stable statethat was based on spatial differences in J. lalandii densities. In our case,wehave demonstrated both spatial (this paper) and temporal (Blamey etal., 2010) differences in lobster abundance EOCH, and we argue that thisecosystem meets the required criteria for a regime shift/alternate stablestate as follows: (1) No equivalent range expansion has been recorded inthe past, so the event can be classed as ‘low frequency’. (2) The responsehas been ‘high amplitude’, with lobsters, urchins, macroalgae andencrusting corallines all changing >46%. (3) The shift was abrupt, withlobsters increasing over a 3–4 year period, and urchins declining within1–2 years (Tarr et al., 1996). (4) There is evidence that the systemincorporated a discontinuous regime shift (Fig. 8). (5) The event haspersisted for at least 15 years, which is longer than the lifespan of almostall participating species,with the possible exception of J. lalandii itself. (6)The coastline in which the shift has occurred spans about 30 km andincludes a total kelp-forest area of ca. 7.5 km2 (Tarr, 1993). (7) At leastthree trophic levels have been affected by the shift: lobsters, herbivores,macroalgae and encrusting corallines.

Connell and Sousa (1983) have argued that putative cases ofalternative stable states should be excluded if they involve humanintervention.While we agree that they should not simply bemaintainedby human effects, we would argue that it is unnecessarily restrictive toeliminate cases caused by human actions, because (1) there is stronginterest in the consequences of human effects and (2) it is virtuallyimpossible to exclude human effects, as manifestations such as climatechange and overfishing are now pervasive.

Taken collectively, we would argue that the area EOCH sufficientlymeets the criteria to consider that the current kelp-forest ecosystemhas shifted into an alternative stable state, maintained by J. lalandii.

4.8. Conclusion

Numbers of J. lalandii have increased EOCH since the early 1990sand they are now significantly more abundant at sites designated as‘invaded’ than theywere in the past (Blamey et al., 2010), and are 20-fold more abundant there than in ‘non-invaded’ sites. In parallel, thebenthic community in invaded areas is significantly different fromthat in non-invaded areas. Our observational and correlative datarevealed strong patterns suggesting that J. lalandii is responsible forthis shift in community structure, although they do not test causeand effect. However, given the broad diet of J. lalandii, the densities ithas reached in invaded areas, and its documented ability to altercommunity structure (Barkai and Branch, 1988a,b; Barkai andMcQuaid, 1988), as well as the observed correlations betweenlobster increases and urchin decreases in the early 1990s (Mayfieldand Branch, 2000; Tarr et al., 1996), the accumulative evidencestrongly suggests that J. lalandii is the cause of the contrastingbenthic communities. We argue that it was responsible for a trophiccascade in invaded areas, depleting herbivores and consequentlyreleasing macroalgae, ultimately leading to a reduction of encrustingcorallines. At non-invaded sites, J. lalandii was rare, herbivoresabundant, macroalgae scarce and encrusting corallines more abun-dant. As depth increased, these effects were still detectable but weredamped-down because (a) differences in abundance of J. lalandiibetween invaded and non-invaded areas were not as obvious, and(b) many of the species were more scarce, reducing their interaction

45L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

strength. Patterns thus reflected the combined effects of J. lalandiiand depth.

The shallows, where the strongest effects of J. lalandii werefelt, coincided with the depth of maximum abundance of urchins(P. angulosus), abalone (H. midae), kelp (E. maxima) and encrustingcorallines, with important ecological and economic repercussions.Firstly, where J. lalandii numberswere high, P. angulosuswas eliminatedand other grazers greatly reduced. Given that P. angulosus provides vitalshelter for juvenile abalone (Day and Branch, 2000a,b, 2002a), itsremoval has imminent negative consequences for the abalone fishery,the heart of which coincides with the area of lobster invasion. Inaddition, the cascading effects of J. lalandii appear to have diminishedencrusting corallines,which are the preferred settlement site of abalonerecruits (Day and Branch, 2000a), adding further stress to thefishery forabalone, which is already in turmoil because of illegal fishing. Theplummeting fishery led to a ban on all fishing for wild abalone stocks(Hauck, 2009) and the rapid development of abalone mariculture(Troell et al., 2006), and the latter has increased the demand forharvested kelp as a food source for cultured abalone. On the positiveside, the increase in J. lalandii has allowed the development of acommercial rock-lobster fishery EOCH. Balancing the harvesting ofthese resources will require deft management and a clear understand-ing of the ecosystem.

Supplementary related to this article can be found online at doi:10.1016/j.jembe.2012.03.022.

Acknowledgments

We gratefully acknowledge all the divers who assistedwith the datacollection; Dr RJ Anderson and Prof CL Griffiths for help with theseaweed and invertebrate identification, and to the Department ofAgriculture, Forestry and Fisheries for access to the FIMS data. Thanks toDrs AC Cockcroft, S Mayfield and ADM Smith for helpful comments onearlier drafts of this manuscript and to two anonymous reviewers fortheir constructive criticism. Funding was provided by the University ofCape Town, the National Research Foundation, and the AndrewMellonFoundation [SS].

References

Anderson, R.J., Carrick, P., Levitt, G.J., Share, A., 1997. Holdfast of adult kelp Eckloniamaxima provide refuges from grazing for recruitment of juvenile kelps. Mar. Ecol.Prog. Ser. 159, 265–273.

Anderson, M.J., Gorley, R.N., Clarke, K.R., 2008. PERMANOVA+ for PRIMER: Guide toSoftware and Statistical Methods. PRIMER-E, Plymouth, UK.

Andrew, N.L., 1993. Spatial heterogeneity, sea urchin grazing, and habitat structure onreefs in temperate Australia. Ecology 74, 292–302.

Andrew, N.L., Underwood, A.J., 1993. Density-dependent foraging in the sea urchinCentrostephanus rodgersii on shallow subtidal reefs in New South Wales, Australia.Mar. Ecol. Prog. Ser. 99, 89–98.

Attwood, C., Farquhar, M., 1999. Collapse of linefish stocks between Cape Hangklip andWalker Bay, South Africa. S. Afr. J. Mar. Sci. 21, 415–432.

Awad, A., Griffiths, C.L., Turpie, J., 2002. Distribution of South African marine benthicinvertebrates applied to the selection of priority conservation areas. Divers. Distrib.8, 129–145.

Babcock, R.C., Kelly, S., Shears, N.T., Walker, J.W., Willis, T.J., 1999. Changes in communitystructure in temperate marine reserves. Mar. Ecol. Prog. Ser. 189, 125–134.

Barkai, A., Branch, G.M., 1988a. Contrasts between the benthic communities of subtidalhard substrata at Marcus and Malgas Islands: a case study of alternative stablestates? S. Afr. J. Mar. Sci. 7, 117–137.

Barkai, A., Branch, G.M., 1988b. The influence of predation and substratal complexityon recruitment to settlement plates: a test of the theory of alternative states. J. Exp.Mar. Biol. Ecol. 124, 215–237.

Barkai, A., Branch, G.M., 1988c. Energy requirements for a dense population of rocklobsters Jasus lalandii: novel importance of unorthodox food sources. Mar. Ecol.Prog. Ser. 50, 83–96.

Barkai, A., McQuaid, C., 1988. Predator–prey role reversal in a marine benthic ecosystem.Science 242, 62–64.

Barkai, A., Davis, C.L., Tugwell, S., 1996. Prey selection by the South African cape rocklobster Jasus lalandii: ecological and physiological approaches. Bull. Mar. Sci. 58, 1–8.

Barrett, N.S., Buxton, C.D., Edgar, G.J., 2009. Changes in invertebrate and macroalgalpopulations in Tasmanian marine reserves in the decade following protection. J.Exp. Mar. Biol. Ecol. 370, 104–119.

Blamey, L.K., 2010. Ecosystem effects of a rock-lobster ‘invasion’: comparative andmodelling approaches. Ph.D. thesis, University of Cape Town, Cape Town.

Blamey, L.K., Branch, G.M., Reaugh-Flower, K.E., 2010. Temporal changes in kelp-forestbenthic communities following an invasion by the rock lobster Jasus lalandii. Afr. J.Mar. Sci. 32, 481–490.

Branch, G.M., 2007. Trophic interactions in subtidal rocky reefs on the west coast ofSouth Africa. In: Branch, G.M., McClanahan, T.R. (Eds.), Food Webs and theDynamics of Marine Reefs. Oxford University Press, Inc., New York, pp. 50–78.

Breen, P.A., Mann, K.H., 1976. Changing lobster abundance and the destruction of kelpbeds by sea urchins. Mar. Biol. 34, 137–142.

Breitburg, D.L., 1984. Residual effects of grazing: inhibition of competitor recruitmentby encrusting coralline algae. Ecology 65, 1136–1143.

Bulleri, F., Bertocci, I., Micheli, F., 2002. Interplay of encrusting coralline algae and seaurchins in maintaining alternative habitats. Mar. Ecol. Prog. Ser. 243, 101–109.

Byrnes, J., Stachowicz, J.J., Hultgren, K.M., Hughes, A.R., Olyarnik, S.V., Thornbert, C.S.,2006. Predator diversity strengthens trophic cascades in kelp forests by modifyingherbivore behaviour. Ecol. Lett. 9, 61–71.

Carpenter, R.C., 1986. Partitioning herbivory and its effects on coral reef algal communities.Ecol. Monogr. 56, 345–364.

Carpenter, R.C., 1988. Mass mortality of Diadema antillarum 1. Long-term effects on seaurchin population-dynamics and coral reef algal communities. Mar. Biol. 104, 67–77.

Castilla, J.C., Moreno, C., 1982. Sea urchins and Macrocystis pyrifera: experimental test ontheir ecological relations in southern Chile. In: Lawrence, J.M. (Ed.), InternationalEchinoderm Conference, Tampa Bay. AA Balkema, Rotterdam, pp. 257–263.

Clark, R.P., Edwards, M.S., Foster, M.S., 2004. Effects of shade from multiple kelpcanopies on an understory algal assemblage. Mar. Ecol. Prog. Ser. 267, 107–119.

Clarke, K.R., Gorley, R.N., 2006. PRIMER v6: User Manual/Tutorial. PRIMER-E, Plymouth,UK.

Cockcroft, A.C., van Zyl, D., Hutchings, L., 2008. Large-scale changes in the spatialdistribution of South African West Coast rock lobsters: an overview. Afr. J. Mar. Sci.30, 149–159.

Collie, J.S., Richardson, K., Steele, J.H., 2004. Regime shifts: can ecological theory illuminatethe mechanisms. Prog. Oceanogr. 60, 281–302.

Connell, J.H., 1978. Diversity in tropical rain forests and coral reefs. Science 199,1302–1310.

Connell, J.H., Sousa, W.P., 1983. On the evidence needed to judge ecological stability orpersistence. Am. Nat. 121, 789–824.

Cowen, R.K., 1983. The effects of sheephead (Semicossyphus pulcher) predation on redsea urchin (Strongylocentrotus franciscanus) populations: an experimental analysis.Oecologia 58, 249–255.

Cox, C., Hunt, J.H., Lyons, W.G., Davis, G.E., 1997. Nocturnal foraging of the Caribbeanspiny lobster (Panulirus argus) on offshore reefs of Florida, USA. Mar. FreshwaterRes. 48, 671–679.

Cury, P., Shannon, L.J., 2004. Regime shifts in upwelling ecosystems: observed changesand possible mechanisms in the northern and southern Benguela. Prog. Oceanogr.60, 223–243.

Cury, P., Bakun, A., Crawford, R.J.M., Jarre, A., Quiñones, R.A., Shannon, L.J.,Verheye, H.M., 2000. Small pelagics in upwelling systems: patterns ofinteraction and structural changes in “wasp-waist” ecosystems. ICES J. Mar.Sci. 57, 603–618.

Day, E.G., Branch, G.M., 2000a. Relationships between recruits of abalone Haliotismidae, encrusting corallines and the sea urchin Parechinus angulosus. S. Afr. J. Mar.Sci. 22, 137–144.

Day, E.G., Branch, G.M., 2000b. Evidence for a positive relationship between juvenileabalone Haliotis midae and the sea urchin Parechinus angulosus in the south-western Cape, South Africa. S. Afr. J. Mar. Sci. 22, 145–156.

Day, E.G., Branch, G.M., 2002a. Effects of sea urchins (Parechinus angulosus) on recruitsand juveniles of abalone (Haliotis midae). Ecol. Monogr. 72, 133–149.

Day, E.G., Branch, G.M., 2002b. Influences of the sea urchin Parechinus angulosus(Leske) on the feeding behaviour and activity rhythms of juveniles of the SouthAfrican abalone Haliotis midae Linn. J. Exp. Mar. Biol. Ecol. 276, 1–17.

Dayton, P.K., 1985a. Ecology of kelp communities. Annu. Rev. Ecol. Syst. 16, 215–245.Dayton, P.K., 1985b. The structure and regulation of some South American kelp

communities. Ecol. Monogr. 55, 447–468.Dayton, P.K., Curie, V., Gerrodette, T., Keller, B., Rosenthal, R., Van Tresca, D., 1984. Patch

dynamics and stability of some southern California kelp communities. Ecol.Monogr. 54, 253–289.

Dayton, P.K., Tegner, M.J., Parnell, P.E., Edwards, P.B., 1992. Temporal and spatialpatterns of disturbance and recovery in a kelp forest community. Ecol. Monogr. 62,421–445.

Dayton, P.K., Tegner, M.J., Edwards, P.B., Riser, K.L., 1998. Sliding baselines, ghosts andreduced expectations in kelp forest communities. Ecol. Appl. 8, 309–322.

DEAT [Department of Environmental Affairs and Tourism], 2005. Policy for theAllocation and Management of Commercial Fishing Rights in the West Coast RockLobster Limited Commercial (nearshore) Fishery: 2005. Marine and CoastalManagement, Department of Environmental Affairs and Tourism, South Africa.

deYoung, B., Harris, R., Alheit, J., Beaugrand, G., Mantua, N., Shannon, L., 2004. Detectingregime shifts in the ocean: data considerations. Prog. Oceanogr. 60, 143–164.

Duggins, D.O., Dethier, M.N., 1985. Experimental studies of herbivory and algalcompetition in a low intertidal habitat. Oecologia 67, 183–191.

Edgar, G.J., 1990. Predator–prey interactions in seagrass beds. 1. The influence ofmacrofaunal abundance and size-structure on the diet and growth of the westernrock lobster Panulirus cygnus George. J. Exp. Mar. Biol. Ecol. 139, 1–22.

Edwards, M.S., 1998. Effects of long-term kelp canopy exclusion on the abundanceof the annual alga Desmarestia ligulata (Light F). J. Exp. Mar. Biol. Ecol. 228,309–326.

46 L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

Estes, J.A., 2007. Kelp forest food webs in the Aleutian Archipelago. In: Branch, G.M.,McClanahan, T.R. (Eds.), Food Webs and the Dynamics of Marine Reefs. OxfordUniversity Press, Inc., New York, pp. 29–49.

Estes, J.A., Duggins, D.O., 1995. Sea otters and kelp forests in Alaska: generality andvariation in a community ecological paradigm. Ecol. Monogr. 65, 75–100.

Estes, J.A., Palmisano, J.F., 1974. Sea otters: their role in structuring nearshore communities.Science 185, 1058–1060.

Estes, J.A., Duggins, D.O., Rathbun, G.B., 1989. The ecology of extinctions in kelp forestcommunities. Conserv. Biol. 3, 252–264.

Estes, J.A., Tinker, M.T., Williams, T.M., Doak, D.F., 1998. Killer whale predation on seaotters linking oceanic and nearshore ecosystems. Science 282, 473–476.

Fairweather, T.P., van der Lingen, C.D., Booth, A.J., Drapeau, L., van der Westhuizen, L.L.,2006. Indicators of sustainable fishing for South African sardine Sardinops sagaxand anchovy Engraulis encrasicolus. Afr. J. Mar. Sci. 28, 661–680.

Fariña, J.M., Palma, A.T., Ojeda, F.P., 2007. Subtidal kelp-associated communities off thetemperate Chilean coast. In: Branch, G.M., McClanahan, T.R. (Eds.), Food Webs andthe Dynamics of Marine Reefs. Oxford University Press, Inc., New York, pp. 79–102.

Field, J.G., Jarman, N.G., Dieckmann, G.S., Griffiths, C.L., Velimirov, B., Zoutendyk, P.,1977. Sun, waves, seaweed and lobsters: the dynamics of a West Coast kelp bed. S.Afr. J. Mar. Sci. 73, 7–10.

Field, J.G., Griffiths, C.L., Griffiths, R.J., Jarman, N., Zoutendyk, P., Velimirov, B., Bowes, A.,1980a. Variation in structure and biomass of kelp communities along the south-west Cape coast. Trans. R. Soc. S. Afr. 44, 145–203.

Field, J.G., Griffiths, C.L., Linley, E.A., Carter, R.A., Zoutendyk, P., 1980b. Upwelling in anearshore marine ecosystem and its biological implications. Estuar. Coast. Mar. Sci.11, 133–150.

Fletcher, W.J., 1987. Interactions among subtidal Australian sea urchins, gastropodsand algae: effects of experimental removals. Ecol. Monogr. 57, 89–109.

Frank, K.T., Petrie, B., Choi, J.S., Leggett, W.C., 2005. Trophic cascades in a formerly cod-dominated ecosystem. Science 308, 1621–1623.

Frank, K.T., Petrie, B., Shackell, N.L., 2007. The ups and downs of trophic control incontinental shelf ecosystems. Trends Ecol. Evol. 22, 236–242.

Frederiksen, M., Edwards, M., Richardson, A.J., Halliday, N.C., Wanless, S., 2006. Fromplankton to top predators: bottom-up control of a marine food web across fourtrophic levels. J. Anim. Ecol. 75, 1259–1268.

Fricke, A.H., 1979. Kelp grazing by the common sea urchin Parechinus angulosus Leskein False Bay, Cape. S. Afr. J. Zool. 14, 143–148.

Gascoigne, J.C., Lipcius, R.N., 2004. Allee effects driven by predation. J. Appl. Ecol. 41,801–810.

Gillett, N.P., Zwiers, F.W., Weaver, A.J., Stott, P.A., 2003. Detection of human influenceon sea-level pressure. Nature 422, 292–294.

Glynn, P.W., 1976. Some physical and biological determinants of coral communitystructure in the eastern Pacific. Ecol. Monogr. 46, 431–456.

Goñi, R., Quetglas, A., Reñones, O., 2001. Diet of the spiny lobster Palinurus elephas(Decapoda: Palinuridea) from the Columbretes Islands Marine Reserve (north-western Mediterranean). J. Mar. Biol. Assoc. UK 81, 347–348.

Götz, A., Kerwath, S.E., Attwood, C.G., Sauer, W.H.H., 2009a. Effects of fishing on atemperate reef community in South Africa 1: ichthyofauna. Afr. J. Mar. Sci. 31,241–251.

Götz, A., Kerwath, S.E., Attwood, C.G., Sauer, W.H.H., 2009b. Effects of fishing on atemperate reef community in South Africa 2: benthic invertebrates and algae. Afr. J.Mar. Sci. 31, 253–262.

Graham, M., Halpern, B., Carr, M., 2007. Diversity and dynamics of Californiansubtidal kelp forests. In: Branch, G.M., McClanahan, T.R. (Eds.), Food Webs andthe Dynamics of Marine Reefs. Oxford University Press, Inc., New York, pp.103–134.

Griffiths, M.H., 2000. Long-term trends in catch and effort of commercial linefish offSouth Africa's Cape Province: snapshots of the 20th century. S. Afr. J. Mar. Sci. 22,81–110.

Grime, J.P., 1973. Competitive exclusion in herbaceous vegetation. Nature 242, 344–347.Guidetti, P., 2006. Marine reserves re-establish lost predatory interactions and cause

community changes in rocky reefs. Ecol. Appl. 16, 967–976.Hagen, N.T., 1995. Recurrent destructive grazing of successionally immature kelp

forests by green sea urchins in Vestfjorden, Northern Norway. Mar. Ecol. Prog. Ser.123, 95–106.

Haley, C., Blamey, L.K., Atkinson, L.J., Branch, G.M., 2011. Dietary change of the rocklobster Jasus lalandii after an ‘invasive’ geographic shift: effects of size, density andfood availability. Estuarine Coastal Shelf Sci. 93, 160–170.

Halpern, B.S., Cottenie, K., Broitman, B.R., 2006. Strong top-down control in southernCalifornia kelp forest ecosystems. Science 312, 1230–1232.

Hauck, M., 2009. Crime, environment and power: revisiting the abalone fishery. S. Afr. J.Crim. Justice 22, 229–245.

Hauck, M., Sweijd, N.A., 1999. A case study of abalone poaching in South Africa and itsimpact on fisheries management. ICES J. Mar. Sci. 56, 1024–1032.

Hay, M.E., 1984. Patterns of fish and urchin grazing on Caribbean coral reefs: areprevious results typical? Ecology 65, 446–454.

Hily, C., Potin, P., Floc'h, J., 1992. Structure of subtidal algal assemblages on soft-bottomsediments: fauna/flora interactions and role of disturbances in the Bay of Brest,France. Mar. Ecol. Prog. Ser. 85, 115–130.

Himmelman, J.H., Cardinal, A., Bourget, E., 1983. Community development followingremoval of urchins, Strongylocentrotus droebachiensis, from the rocky subtidal zoneof the St. Lawrence Estuary, eastern Canada. Oecologia 59, 27–39.

Huston, M.A., 1985a. Patterns of species diversity on coral reefs. Annu. Rev. Ecol. Syst.16, 149–177.

Huston, M.A., 1985b. Patterns of species diversity in relation to depth at Discovery Bay,Jamaica. Bull. Mar. Sci. 37, 928–935.

Huston, M.A., 1994. Biological Diversity: The Coexistence of Species on ChangingLandscapes. Cambridge University Press, Cambridge, UK.

Jackson, J.B.C., Kirby, M.X., Berger, W.H., Bjorndal, K.A., Botsford, L.W., Bourque,B.J., Bradbury, R.H., Cooke, R., Erlandson, J., Estes, J.A., Hughes, T.P., Kidwell, S.,Lange, C.B., Lenihan, H.S., Pandolfi, J.M., Peterson, C.H., Steneck, R.S., Tegner, M.J.,Warner, R.R., 2001. Historical overfishing and the recent collapse of coastalecosystems. Science 293, 629–638.

Johnson, C.R., Mann, K.H., 1986. The crustose coralline alga, Phymatolithon foslie,inhibits the overgrowth of seaweeds without relying on herbivores. J. Exp. Mar.Biol. Ecol. 96, 127–146.

Keats, D.W., Wilton, P., Maneveldt, G., 1994. Ecological significance of deep-layersloughing in the eulittoral zone coralline alga, Spongites yendoi (Foslie) Chamberlain(Corallinaceae, Rhodophyta) in South Africa. J. Exp. Mar. Biol. Ecol. 175, 145–154.

Keats, D.W., Knight, M.A., Pueschel, C.M., 1997. Anti-fouling effects of epithallialshedding in three crustose coralline algae (Rhodophyta, Corallinales) on a coralreef. J. Exp. Mar. Biol. Ecol. 213, 281–293.

Kennelly, S.J., 1989. Effects of kelp canopies on understory species due to shade andscour. Mar. Ecol. Prog. Ser. 50, 215–224.

Köster, F.W., Möllmann, C., 2000. Trophodynamic control by clupeid predators onrecruitment success in Baltic cod. ICES J. Mar. Sci. 57, 310–323.

Lang, C., Mann, K.H., 1976. Changes in sea urchin populations after the destruction ofkelp beds. Mar. Biol. 36, 321–326.

Lawrence, J.M., 1975. On the relationships between marine plants and sea urchins.Oceanogr. Mar. Biol. Annu. Rev. 13, 213–286.

Lees, K., Pitois, S., Scott, C., Frid, C., Mackinson, S., 2006. Characterizing regime shifts inthe marine environment. Fish Fish. 7, 104–127.

Ling, S.D., 2008. Range expansion of a habitat-modifying species leads to loss oftaxonomic diversity: a new and impoverished reef state. Oecologia 156, 883–894.

Ling, S.D., Johnson, C.R., 2009. Population dynamics of an ecologically important range-extender: kelp-beds versus sea urchin barrens. Mar. Ecol. Prog. Ser. 374, 113–125.

Ling, S.D., Johnson, C.R., Frusher, S., King, C.K., 2008. Reproductive potential of a marineecosystem engineer at the edge of a newly expanded range. Global Change Biol. 14,1–9.

Ling, S.D., Johnson, C.R., Ridgway, K., Hobday, A.J., Haddon, M., 2009a. Climate-drivenrange extension of a sea urchin: inferring future trends by analysis of recentpopulation dynamics. Global Change Biol. 15, 719–731.

Ling, S.D., Johnson, C.R., Frusher, S., Ridgway, K., 2009b. Overfishing reduces resilienceof kelp beds to climate-driven catastrophic phase shift. Proc. Natl. Acad. Sci. U. S. A.106, 22341–22345.

Littler, M.M., Littler, D.S., Taylor, P.R., 1995. Selective herbivore increases biomass of itsprey: a chiton-coralline reef-building association. Ecology 76, 1666–1681.

Loya, Y., 1972. Community structure and species diversity of hermatypic corals at Eilat,Red Sea. Mar. Biol. 13, 100–123.

Mann, K.H., 1977. Destruction of kelp-beds by sea urchins: a cyclical phenomenon orirreversible degradation? Helgol. Wiss. Meeres. 30, 455–467.

Mann, K.H., Breen, P.A., 1972. The relation between lobster abundance, sea urchins and,kelp beds. J. Fish. Res. Board Can. 29, 603–609.

Mayfield, S., Branch, G.M., 2000. Interrelations among rock lobsters, sea urchins andjuvenile abalone: implications for community management. Can. J. Fish. Aquat. Sci.57, 2175–2185.

Mayfield, S., Atkinson, L.J., Branch, G.M., Cockcroft, A.C., 2000a. Diet of the West Coastrock lobster Jasus lalandii: influence of lobster size, sex, capture depth, latitude andmoult stage. S. Afr. J. Mar. Sci. 22, 57–69.

Mayfield, S., Branch, G.M., Cockcroft, A.C., 2000b. Relationships among diet, growthrate, and food availability for the South African rock lobster, Jasus lalandii(Decapoda, Palinuridea). Crustaceana 73, 815–834.

Mayfield, S., De Beer, E., Branch, G.M., 2001. Prey preference and the consumption ofsea urchins and juvenile abalone by captive rock lobsters (Jasus lalandii). Mar.Freshwater Res. 52, 773–780.

Mead, A., 2011 Climate and bioinvasions as drivers of change on South African rockyshores. Ph.D thesis, University of Cape Town, Cape Town.

Melville, A.J., Connell, S.D., 2001. Experimental effects of kelp canopies on subtidalcoralline algae. Austral Ecol. 26, 102–108.

Menge, B.A., Lubchenco, J., 1981. Community organization in temperate and tropicalrocky intertidal habitats: prey refuges in relation to consumer pressure gradients.Ecol. Monogr. 51, 429–450.

Menge, B.A., Sutherland, J.P., 1976. Species diversity gradients: synthesis of the roles ofpredation, competition, and temporal heterogeneity. Am. Nat. 110, 351–369.

Morrison, D., 1988. Comparing fish and urchin grazing in shallow and deeper coral reefalgal communities. Ecology 69, 1367–1382.

Paine, R.T., 1966. Food web complexity and species diversity. Am. Nat. 100, 65–75.Paine, R.T., 1974. Intertidal community structure. Oecologia 15, 93–120.Paine, R.T., 1980. Food webs: linkage, interaction strength and community infrastruc-

ture. J. Anim. Ecol. 49, 667–685.Paine, R.T., 1984. Ecological determinism in the competition for space. Ecology 65,

1339–1348.Quinn, G.P., Keough, M.J., 2002. Experimental Design and Data Analysis for Biologists.

Cambridge University Press, Cambridge, UK.Quinn, J.F., Wing, S.R., Botsford, L.W., 1993. Harvest refugia in marine invertebrate

fisheries: models and applications to the Red Sea urchin, Strongylocentrotusfranciscanus. Am. Zool. 33, 537–550.

Reed, D.C., Foster, M.S., 1984. The effects of canopy shadings on algal recruitment andgrowth in a giant kelp forest. Ecology 65, 937–948.

Reynolds, C.S., Padisák, J., Sommer, U., 1993. Intermediate disturbance in the ecology ofphytoplankton and themaintenance of species-diversity— a synthesis. Hydrobiologia249, 183–188.

47L.K. Blamey, G.M. Branch / Journal of Experimental Marine Biology and Ecology 420–421 (2012) 33–47

Rice, J., 1995. Food web theory, marine food webs, and what climate change may do tonorthern marine fish populations. In: Beamish, R.J. (Ed.), Climate Change andNorthern Fish Populations: Can. Spec. Publ. Fish. Aquat. Sci., 121, pp. 516–568.

Richardson, A.J., Schoeman, D.S., 2004. Climate impact on plankton ecosystems in thenortheast Atlantic. Science 305, 1609–1612.

Ridgway, K.R., 2007. Long-term trend and decadal variability of the southwardpenetration of the East Australian Current. Geophys. Res. Lett. 34, L13616.doi:10.1029/2007GL030393.

Robles, C., 1987. Predator foraging characteristics and prey population structure on asheltered shore. Ecology 68, 1502–1514.

Robles, C., Sweetnam, D., Eminike, J., 1990. Lobster predation on mussels: shore-leveldifferences in prey vulnerability and predator preference. Ecology 71, 1564–1577.

Rouault, M., Penven, P., Pohl, B., 2009. Warming in the Agulhas Current system sincethe 1980s. Geophys. Res. Lett. 36, L12602. doi:10.1029/2009GL037987.

Roy, C., van der Lingen, C.D., Coetzee, J.C., Lutjeharms, J.R.E., 2007. Abrupt environmentalshift associated with changes in the distribution of Cape anchovy Engraulisencrasicolus spawners in the southern Benguela. Afr. J. Mar. Sci. 29, 309–319.

Sala, E., Zabala, M., 1996. Fish predation and the structure of the sea urchin Paracentrotuslividus populations in the NW Mediterranean. Mar. Ecol. Prog. Ser. 140, 71–81.

Sala, E., Boudouresque, C.F., Harmelin-Vivien, M., 1998. Fishing, trophic cascades, andthe structure of algal assemblages: evaluation of an old but untested paradigm.Oikos 82, 425–439.

Santelices, B., Ojeda, F.P., 1984. Population dynamics of coastal forests of Macrocystispyrifera in Peurto Tora, Isla Navarino, Southern Chile. Mar. Ecol. Prog. Ser. 14, 175–183.

Scheffer, M., Carpenter, S.R., 2003. Catastrophic regime shifts in ecosystems: linkingtheory to observation. Trends Ecol. Evol. 18, 648–656.

Seidel, D.J., Fu, Q., Randel, W.J., Reichler, T.J., 2008. Widening of the tropical belt in achanging climate. Nat. Geosci. 1, 21–24.

Shannon, L.V., 1985. The Benguela ecosystem. Part I. Evolution of the Benguela, physicalfeatures and processes. Oceanogr. Mar. Biol. Annu. Rev. 23, 105–182.

Shea, K., Roxburgh, S.H., Rauschert, E.S.J., 2004. Moving from pattern to process:coexistence mechanisms under intermediate disturbance regimes. Ecol. Lett. 7,491–508.

Shears, N.T., Babcock, R.C., 2002. Marine reserves demonstrate top-down control ofcommunity structure on temperate reefs. Oecologia 132, 131–142.

Shears, N.T., Babcock, R.C., 2003. Continuing trophic cascade effects after 25 years ofno-take marine reserve protection. Mar. Ecol. Prog. Ser. 246, 1–16.

Sousa, W.P., 1979. Disturbance in marine intertidal boulder fields: the nonequilibriummaintenance of species diversity. Ecology 60, 1225–1239.

Steneck, R.S., 1982. A limpet-coralline alga association: adaptations and defencesbetween a selective herbivore and its prey. Ecology 63, 507–522.

Steneck, R.S., 1986. The ecology of coralline algal crusts: consequent patterns andadaptive strategies. Annu. Rev. Ecol. Syst. 17, 273–303.

Steneck, R.S., 1998. Human influences on coastal ecosystems: does overfishing createtrophic cascades? Trends Ecol. Evol. 13, 429–430.

Steneck, R.S., Dethier, M.N., 1994. A functional group approach to the structure of algal-dominated communities. Oikos 69, 476–498.

Steneck, R.S., Graham, M.H., Bourque, B.J., Corbett, D., Erlandson, J.M., Estes, J.A., Tegner,M.J., 2002. Kelp forest ecosystems: biodiversity, stability, resilience and future.Environ. Conserv. 29, 436–459.

Steneck, R.S., Vavrinec, J., Leland, A.V., 2004. Accelerating trophic-level dysfunction inkelp forest ecosystems of the western north Atlantic. Ecosystems 7, 323–332.

Strong, D.R., 1992. Are trophic cascades all wet? Differentiation and donor-control inspeciose ecosystems. Ecology 73, 747–754.

Swain, D.P., Sinclair, A.F., 2000. Pelagic fishes and the cod recruitment dilemma in theNorthwest Atlantic. Can. J. Fish. Aquat. Sci. 57, 1321–1325.

Tarr, R.J.Q., 1993. Stock assessment, and aspects of the biology of the South Africanabalone, Haliotis midae. M.Sc. Thesis, University of Cape Town, Cape Town.

Tarr, R.J.Q., Williams, P.V.G., Mackenzie, A.J., 1996. Abalone, sea urchins and rocklobster: a possible ecological shift that may affect traditional fisheries. S. Afr. J. Mar.Sci. 17, 319–323.

Tegner, M.J., Dayton, P.K., 1981. Population structure, recruitment and mortality of twosea urchins (Strongylocentrotus franciscanus and S. purpuratus) in a kelp forest.Mar. Ecol. Prog. Ser. 5, 255–268.

Tegner, M.J., Dayton, P.K., 1991. Sea urchins, El Niños and the long term stability ofSouthern California kelp forest communities. Mar. Ecol. Prog. Ser. 77, 49–63.

Tegner, M.J., Dayton, P.K., 2000. Ecosystem effects of fishing in kelp forest communities.ICES J. Mar. Sci. 57, 579–589.

Tegner, M.J., Levin, L.A., 1983. Spiny lobsters and sea urchins: analysis of a predator–prey interaction. J. Exp. Mar. Biol. Ecol. 73, 125–150.

Thompson, D.W.J., Solomon, S., 2002. Interpretation of recent Southern Hemisphereclimate change. Science 296, 895–899.

Troell, M., Robertson-Andersson, D., Anderson, R.J., Bolton, J.J., Maneveldt, G., Halling,C., Probyn, T., 2006. Abalone farming in South Africa: An overview withperspectives on kelp resources, abalone feed, potential for on-farm seaweedproduction and socio-economic importance. Aquaculture 257, 266–281.

Turpie, J.K., Beckley, L.E., Katua, S.M., 2000. Biogeography and the selection of priorityareas for conservation of South African coastal fishes. Biol. Conserv. 92, 59–72.

Vance, R.R., 1979. Effects of grazing by the sea urchin, Centrostephanus coronatus, onprey community composition. Ecology 60, 537–546.

Vanderklift, M.A., Kendrick, G.A., 2005. Contrasting influence of sea urchins on attachedand drift macroalgae. Mar. Ecol. Prog. Ser. 299, 101–110.

Velimirov, B., Griffiths, C.L., 1979. Wave induced kelp movement and its importance forcommunity structure. Bot. Mar. 22, 169–172.

Velimirov, B., Field, J.G., Griffiths, C.L., Zoutendyk, P., 1977. The ecology of kelp bedcommunities in the Benguela upwelling system. Helgol. Wiss. Meeres. 30, 495–518.

Ware, D.M., Thompson, R.E., 2005. Bottom-up ecosystem dynamics determine fishproduction in the northeast Pacific. Science 308, 1280–1284.

Watanabe, J.M., Harrold, C., 1991. Destructive grazing by sea urchins Strongylocentrotusspp. in a central California kelp forest: potential roles of recruitment, depth andpredation. Mar. Ecol. Prog. Ser. 71, 125–141.

Watson, J., Estes, J.A., 2011. Stability, resilience, and phase shifts in rocky subtidalcommunities along the west coast of Vancouver Island, Canada. Ecol. Monogr. 81,215–239.

Wootton, J.T., 1995. Effects of birds on sea urchins and algae: a lower-intertidal trophiccascade. Ecoscience 2, 321–328.

Worm, B., Myers, R.A., 2003. Meta-analysis of cod-shrimp interactions reveals top-down control in oceanic food webs. Ecology 84, 162–173.