Echinoid community structure and rates of herbivory and bioerosion

Transcription

Journal of Experimental Marine Biology and Ecology 456 (2014) 8–17
Contents lists available at ScienceDirect
Journal of Experimental Marine Biology and Ecology
journal homepage: www.elsevier.com/locate/jembe
Echinoid community structure and rates of herbivory and bioerosion on
exposed and sheltered reefs
Omri Bronstein ⁎, Yossi Loya
Department of Zoology, The George S. Wise Faculty of Life Sciences, Tel-Aviv University, Tel-Aviv 69978, Israel
a r t i c l e
i n f o
Article history:
Received 19 September 2013
Received in revised form 24 January 2014
Accepted 9 March 2014
Available online xxxx
Keywords:
Bioerosion
Coral reefs
Herbivory
MPAs
Sea urchins
Western Indian Ocean
a b s t r a c t
Echinoid–habitat relations are complex and bi-directional. Echinoid community structure is affected by the habitat structural and environmental conditions; while at the same time, echinoids may also act as ‘reef engineers’,
able to alter marine environments on a wide geographic scale. In particular, echinoids play a major role in
bioerosion and herbivory on coral reefs. Through feeding, echinoids reduce algal cover, enabling settlement
and coral growth. However, at the same time, they also remove large parts of the reef hard substrata, gradually
leading to reef degradation. Here, we compared coral and macroalgal abundance, echinoid community structure
and species-specific rates of echinoid herbivory and bioerosion on reefs subjected to different intensities of oceanic exposure. Spatio-temporal variations in coral and macroalgal cover were monitored, and populations of the
four most abundant echinoid species on the coral reefs of Zanzibar – Diadema setosum (Leske), Diadema savignyi
(Michelin), Echinometra mathaei (de Blainville) and Echinothrix diadema (Linnaeus) – were compared between
the Island's eastern exposed reefs and western sheltered ones. To account for the effect of management in the
context of reef exposure, we included marine protected areas (MPAs) of both types of reef categories (i.e. sheltered and exposed) in our comparison. Coral and macroalgal cover presented a conspicuous contrasting pattern
across exposed and sheltered sites. While coral dominance and lack of macroalgae were prominent on sheltered
reefs, an opposite trend of low coral cover and moderate–high macroalgal cover were found on exposed reefs.
Bioerosion was also significantly higher on exposed reefs than on sheltered ones (4.2–13 and 1.2–3.9 kg CaCO3
m−2 year−1, respectively). The highest rates, recorded on Pongwe, with almost 7 kg CaCO3 m−2 year−1, are
among the highest echinoid bioerosion rates known to date. Management had a substantial effect on habitat
and echinoid community structure, as coral cover was significantly higher, macroalgal cover lower, and echinoid
densities generally reduced on MPAs regardless of exposure intensity. Our findings suggest that exposed reefs are
susceptible to markedly higher degrees of echinoid bioerosion; however, adequate management measures can
significantly reduce these rates, consequently altering the reef's trajectory for degradation.
© 2014 Elsevier B.V. All rights reserved.
1. Introduction
Common coral-reef associated echinoids have a range of different
feeding modes. Echinoids are considered to be generalist herbivores as
their diets may include algae and seaweed (Klumpp et al., 1993;
Lawrence, 1975; Vaïtilingon et al., 2003), or omnivores due to the inclusion of animal tissue (Briscoe and Sebens, 1988; McClintock et al., 1982),
and even the occasional predation of live coral tissue (Bak and van Eys,
1975; Carpenter, 1981; Glynn et al., 1979). This dietary flexibility,
coupled with their great abundance on some coral reefs (Bauer, 1980;
McClanahan and Kurtis, 1991), place echinoids as keystone species in
coral reef environments. As hard-substrate eroders (Bak, 1990; Glynn
et al., 1979; Hunter, 1977; Trudgill et al., 1987) they scrape the surface
while grazing (Lawrence and Sammarco, 1982), reducing algal cover
(Mapstone et al., 2007) and breaking down reef substratum (Bak,
⁎ Corresponding author. Tel.: +972 3 6409809; fax: +972 3 6727746.
E-mail address: omribronstein@gmail.com (O. Bronstein).
http://dx.doi.org/10.1016/j.jembe.2014.03.003
0022-0981/© 2014 Elsevier B.V. All rights reserved.
1990; Hawkins and Lewis, 1982). At moderate sea urchin densities
this action may facilitate a topographic complexity that favors increased
biodiversity (Johnson et al., 2003) and may also enhance coral recruitment (Birkeland and Randall, 1981; Carpenter and Edmunds, 2006;
Griffin et al., 2003). However, at high sea urchin densities, echinoids
may limit reef growth through predation of coral tissue (Glynn et al.,
1979) or extensive coral (Bak et al., 1984; Mokady et al., 1996) and crustose coralline algae (CCA) erosion (O'leary and McClanahan, 2010).
Moreover, the indiscriminate nature of echinoid grazing has a profound
effect on coral community composition through its control of newlysettled coral spat (Sammarco, 1980, 1982). Consequently, high sea urchin abundance may alter the structure of coral reef communities by
eroding the reef's coral framework, leading to gradual reef degradation.
Many variables have been recognized as important in regulating
echinoid food consumption. For example, species composition, body
size, population densities (Bak, 1990, 1994; Carreiro-Silva and
McClanahan, 2001; Scoffin et al., 1980), attraction to food (Vadas and
Elner, 2003), hydrodynamics (Siddon and Witman, 2003), light (Mills
O. Bronstein, Y. Loya / Journal of Experimental Marine Biology and Ecology 456 (2014) 8–17
et al., 2000; Vaïtilingon et al., 2003), temperature (Larson et al., 1980),
and reproductive stage (Klinger et al., 1997), have all been mentioned
as factors influencing echinoid feeding rates and ecological impact. However, beyond the physiological aspects determined by the life histories of
particular species, echinoid food consumption, and consequently the
rates of herbivory and bioerosion, must be considered in terms of the environmental conditions that exist in their habitats, as gradients in the
physical environment may produce variability in the abundance and distribution of echinoid populations (Andrew, 1993; Clemente and
Hernández, 2008). Several studies have investigated the relationship between coral reef associated echinoids and their habitat (e.g., Dumas et al.,
2007; Graham and Nash, 2013; McClanahan, 1998; McClanahan and
Kurtis, 1991; O'leary and McClanahan, 2010; Peyrot-Clausade et al.,
2000). These publications suggest aspects such as structural complexity,
macroalgal and coral cover, sedimentation, and the presence or absence
of predators, as having substantial effects on the composition, distribution, and size of related echinoid populations. For example, marine
protected areas (MPAs) protecting various echinoid predators consequently present lower rates of sea urchins compared to reefs with
depauperate predatory populations (McClanahan and Kurtis, 1991;
McClanahan et al., 1999). Additionally, echinoid communities tend to
display strong differences in species distribution between exposed and
sheltered reefs, making sea urchin ecology further complex (Dumas
et al., 2007).
Zanzibar Island (Unguja, Tanzania) is situated on the continental
shelf of Tanzania between 50°40′ and 60°30′ south of the equator,
35 km from the mainland. Being an island surrounded by coral reefs, exposed to strong easterly winds and with a sheltered west coast, makes
Zanzibar an ideal study location for echinoid ecology. Located off the
East-African shoreline, the island's coral reefs are fundamental to the
entire marine environment and of great economic importance for the
large human population that depends on them for a livelihood
(Jiddawi, 1997; Khatib, 1997; Mbije et al., 2002; Ngoile and Horrill,
1993). Small patches of mangrove forest and shallow patches of fringing
reefs occur along the more sheltered western coast, while on the more
exposed eastern coast fringing reefs slope up to a narrow coastal lagoon
backed by sand beaches or fossil coral cliffs (Richmond, 2002). The eastern and western sides of the island are subject to markedly different
wave and current intensities; reefs on the eastern ocean-facing side
are exposed to the Indian Ocean (IO) and are susceptible to strong
waves and currents, while reefs in the Zanzibar channel, on the Island's
western side, are sheltered from direct exposure to the IO (Bergman and
Öhman, 2001; Ngoile, 1990). Swell waves generated in the IO can travel
undisturbed for thousands of miles before hitting the Island's eastern
reefs. These swell waves occur off the east coast of Zanzibar for much
of the year, changing their orientation from north-east (between October and March) to south-east (between March and October) depending
on monsoonal season (McClanahan, 1988b; Zanzibar Department of
Environment and MACEMP, 2009). In contrast to the north-east monsoon, the south-east monsoon is characterized by high cloud cover,
rain, high wind energy, decreased temperatures and light, and rougher
seas, with velocities of the East African Coastal Current (EACC) increasing to a speed of four knots (McClanahan, 1988b). The semi-diurnal
tides have mean spring amplitude of 3.3 m, with associated tidal currents being stronger on the east coast, where currents up to three
knots are common (Bergman and Öhman, 2001).
Here, we studied coral and macroalgal cover, echinoid community
structure and associated rates of herbivory and bioerosion on exposed
and sheltered coral reefs. The following hypotheses were tested:
(1) Coral and macroalgal cover vary between exposed and sheltered
reefs. (2) Echinoid community structure, and consequently the intensity
of echinoid-induced bioerosion, are influenced by the degree of oceanic
exposure (e.g., the exposure to higher intensities of waves, currents,
tides, etc.). (3) Rates of echinoid herbivory and bioerosion on marineprotected areas are lower than on unprotected sites. Finally, we present
data on spatio-temporal variations of coral and macroalgal cover, and a
9
detailed account on echinoid community structure and associated rates
of herbivory and bioerosion around the Island of Zanzibar, WIO.
2. Methods
2.1. Study sites
Coral communities and associated echinoid populations were
studied on six reefs surrounding Zanzibar Island (Fig. 1). The sites
were selected to represent sheltered and exposed reefs in terms of oceanic exposure. To test for effects of marine protected areas, MPAs from
both exposure categories (i.e. sheltered and exposed) were selected.
However due to the scarcity of MPAs in the region, only one such site
per exposure category was available for this analysis. Three sites,
Bawe (06°08.7′S; 039°08.2′E), Changu (06°06.8′S; 039°09.8′E), and
Chumbe (06°16.3′S; 039°10.2′E), were selected on the sheltered
western side of the main island facing the Zanzibar channel. The site
at Changu is located ca. 5.5 km from Zanzibar Town and a similar
distance from the site at Bawe. Chumbe is located ca. 12 km south
of Zanzibar Town, and has been a private nature reserve, developed
and managed by the Chumbe Island Coral Park (CHICOP), since
1992 (Nordlund and Walther, 2010). The sites on the exposed eastern
side of Zanzibar were Kiwengwa (06°00.9′S; 039°24.6′E), Pongwe
(06°01.9′S; 039°25.2′E), and Mnemba (05°48.5′S; 039°21.3′E). The
N
AFRICA
Mnemba
Kiwengwa
ZANZIBAR
Pongwe
Bawe
Changu
Chumbe
10 KM
Fig. 1. Map of Zanzibar showing the six study sites. Double circles indicate sites are
marine-protected areas.
10
former two sites are located ca. 3 km from one another on the reef-flats
of Zanzibar's eastern fringing reefs. Mnemba is located ca. 2.5 km off
Zanzibar's north-eastern tip, and has been protected from extractive resource use since 1989 (Bergman and Öhman, 2001).
2.2. Coral and macroalgal cover
Three surveys were conducted between 2006 and 2008 (November
2006, March 2007, and April 2008). At each site, 12 20 × 1 m belttransects were randomly placed to sample the composition of coral assemblages at depths of 3–5 m using SCUBA. Each belt-transect was
subdivided into 20 1 × 1 m quadrats, which were photographed with
a digital camera (Olympus C-5060) attached to a 1 × 1 m PVC-frame.
The percentage coral and macroalgal cover was calculated from the
photographs taken in the field.
2.3. Echinoid community structure
Species assemblages and densities were quantified in March 2008.
Ten randomly placed belt-transects, measuring 20 × 0.5 m, were conducted at each site, at depths of 1–5 m using SCUBA. All echinoids within a transect were recorded and identified to the species level. When
present, the size frequency distributions (SFD) of the dominant species
(i.e. Diadema setosum, Diadema savignyi, Echinometra mathaei and
Echinothrix diadema) from each site were estimated. Estimations were
based on length measurements of the first 100 individuals encountered
in randomly placed one square meter quadrates. The length was measured underwater to the nearest 0.5 mm as the longest axis at the
ambitus, using thin blade Vernier calipers. The mean size of individuals
in the population was then calculated for each species at each site.
2.4. Echinoid gut contents analysis
Individual rates of bioerosion and herbivory of the dominant sea urchin species at each site were determined in experiments conducted at
the Institute of Marine Sciences (IMS) in Zanzibar City during March
2008. Throughout this study, the term ‘bioerosion’ refers to the total
amount of newly eroded CaCO3 from the hard reef substrate, which is
largely composed of scleractinian corals and crustose coralline algae
(CCA). These rates, together with species densities, were then used to
estimate annual echinoid bioerosion and herbivory rates per square
meter for each species at each site. The total bioerosion and herbivory
rates for the entire Zanzibar region were evaluated by pooling all sites
together.
Ten individuals of each species from each site were collected from
the field and brought to the lab. Sampled individuals were chosen according to the mean sizes of each species at the different study sites,
as determined by the size frequency distribution (SFD) estimations
noted above. Collection was carried out on the reef-flats, in early morning, at a depth of 3 m, using SCUBA diving. Individual sea urchins were
separately placed in sealed containers underwater to avoid loss of material during transfer. At the lab, the sea urchins were measured and
weighed after being blotted for five minutes on filter paper. They were
then dissected under a binocular to extract the gut wall from the gut
contents. Extractions were followed by repeated rinses with distilled
water. The gut contents were then analyzed in terms of organic and inorganic fractions, with the latter being further separated into calcium
carbonate (CaCO3) and non-soluble residue fractions (e.g. quartz grains,
sponge spicules, and silt). Analysis of the organic fraction followed a
modification of the ignition-loss method (Dean, 1974). The total extracted gut content was dried in a preheated oven (WTB Binder 1505)
at 60 °C for 48 h (or until a stable weight was reached) and weighed
with a Shimadzu AW220 analytical balance to the nearest 0.0001 g.
The low drying temperature of 60 °C was necessary to minimize the
loss of volatile, especially lipoid, constituents. The samples were then
transferred to a furnace (Carbolite 1200 °C Ashing-plus furnace) and
burned at 550 °C for 6 h in order to eliminate the organic material.
After drying, the samples were weighed again. The difference in weights
before and after ashing was used as a measure of the organic matter in
the sea urchin's gut. The CaCO3 fraction was then determined by acidic
digestion of the residual material after removal of the organic matter.
1.13 N hydrochloric acid (HCl) was used to dissolve the CaCO3. 0.5 g
subsamples of residual ashed material were incubated for 10 min with
25 ml 1.13 N HCl. Following incubation and complete dissolution of
CaCO3, the solution of HCl and residual material was filtered on
0.22 μm PTFE filters (Millipore Hydrophilic Durapore Membrane disk
filters), using a suction filter system. Prior to filtration the filters were
dried in a pre-heated oven at 60 °C for 24 h, and weighed to obtain
the filter's dry weight. Following filtration the filters were dried as before and reweighed. The weight of the residual material retained in
the filter was calculated by subtracting the filter's dry weight before filtration from the dry weight after filtration. This weight corresponds to
the non-soluble residue fraction. Subtracting the weight of the nonsoluble residue fraction from the total ashed weight resulted in the
weight of CaCO3 in the gut. The results obtained in the analysis are
expressed in terms of percentage of each fraction in the sea urchin gut
for each species at each site.
2.5. Estimating the rates of bioerosion and herbivory
The daily food consumption could be estimated based on two
parameters: the average amount of food in the gut, and the number of
hours necessary for complete gastric evacuation (Bajkov, 1935; Elliott,
1972; Elliott and Persson, 1978). The average amount of food in the
gut was calculated for each species at each site as previously described.
The rates of gastric evacuation for the echinoid species studied were obtained from Carreiro-Silva and McClanahan (2001). To calculate the
daily ingestion rates, the average amount of food in the gut and the
rates of gastric evacuation were used in an equation developed by
Elliot and Persson (1978). Assuming an exponential rate of gastric evacuation and a constant rate of food consumption, the daily rates of food
consumption could be calculated using the equation:
F ¼ CR
where the daily food consumption (F) could be estimated from the average amount of food (C) in the stomach at the time of sampling, and the
rate of gastric evacuation (R) (see Carreiro-Silva and McClanahan,
2001).
The two assumptions at the basis of this model – an exponential rate
of gastric evacuation and a constant rate of food consumption – were previously validated for the echinoid species of the current study. An exponential rate of gastric evacuation was demonstrated in all species of the
current study by conducting gut emptying experiments (Carreiro-Silva
and McClanahan, 2001; McClanahan and Kurtis, 1991; Mokady et al.,
1996). The assumption of a constant rate of food consumption is widely
accepted (e.g. McClanahan and Kurtis, 1991; Mokady et al., 1996), and
is supported by field observations (Carreiro-Silva and McClanahan,
2001; Glynn et al., 1979; Klumpp et al., 1993) as well as controlled field
experiments (Downing and El-Zahr, 1987), and is in agreement with
our own observations of sea urchins actively feeding during all hours of
the day.
To estimate the true scale of reef degradation (i.e. the scraping off of
new material from the reef's hard substratum), the source of CaCO3
found in the sea urchins' guts must be considered (Scoffin et al.,
1980). It is therefore essential to distinguish between reworked sediment (i.e. recycling of previously eroded sediment) and newly-eroded
sediment (Bak, 1990; Hunter, 1977; Mokady et al., 1996). The measure
of reworked sediment can then be subtracted from the total amount of
CaCO3 in the gut, and the remaining portion considered as a direct representation of the reef's framework erosion. However, to date there is
little consensus over the way to adequately estimate the amounts of
echinoid sediment consumption (i.e. reworked sediment). Mainly, it
has not been concluded whether sediment ingestion by echinoids is a
process driven by behavior, or merely an outcome of environmental
conditions (e.g., sediment load). Our estimates of the proportion of
reworked sediment and gut turnover rates for D. setosum, D. savignyi,
and E. diadema are based on Carreiro-Silva and McClanahan (2001),
and for E. mathaei from McClanahan and Kurtis (1991). These estimates
represent the highest values of reworked sediment for these species
available from the literature and as such would yield the most conservative (i.e. low) bioerosion estimations. Still, as the impact of sedimentation on echinoid sediment consumption could not be elucidated at
this point, these values should be treated with caution.
11
performed. Permutations were performed using the lmPerm package
(Wheeler, 2010) for data analysis, allowing all permutations of Y (i.e.
Perm = “Exact”). Temporal variations in coral and macroalgal cover
were tested using permutation analysis of variance (pANOVA), and spatial differences (i.e. between sites and sea facing sides) using a nested
design pANOVA with year b site b side. Differences in sea urchin densities were tested using a two-way pANOVA with sites and species as factors. Size frequency distributions were compared using pair-wise
Kolmogorov–Smirnov tests and adjusted for multiple resting using the
Bonferroni correction to minimize false-discovery-rate. Gut content
fractionation was tested using one-way pANOVA. The Tukey Honest Significant Difference (HSD) method which controls for the Type I error
rate across multiple comparisons was used when appropriate.
2.6. Statistical analysis
3. Results
Data analyses were performed using R software for statistical computing (Team, 2013). All data were tested for normality and homogeneity of variance prior to deciding upon the appropriate statistical test. As
data violated test assumptions of normal distribution and homoscedasticity, and as data transformations failed to bring the data to meet the
assumptions of parametric statistical tests, permutation analysis was
Western sites presented significantly higher coral cover in comparison to eastern sites (pANOVA, p b 0.01; Fig. 2A). Throughout the years
of the surveys trends of coral cover have remained consistent within
Coral cover
A
80
70
% coral cover (AVG ± SE)
3.1. Coral community structure
a
W
Nov- 06
E
a a
Mar- 07
Apr- 08
60
50
a
a
a
a
40
a
a
30
20
a
ab
10
b
a a
a a a
0
Changu
Bawe
Chumbe
Mnemba
Kiwengwa
a
Pongwe
Site
Macroalgae cover
B
% macroalgae cover (AVG ± SE)
70
W
E
a
a
Nov- 06
Mar- 07
60
Apr- 08
50
a
c
40
30
20
b
b
10
0
a a a
Changu
a a a
a a a
Bawe
Chumbe
a ab
Mnemba
a
Kiwengwa
Pongwe
Site
Fig. 2. Coral (A) and macroalgal (B) percent cover (mean ± SE) at the different study sites between November 2006 and March 2008. Western sites and eastern sites are denoted W and E,
respectively. Lowercase letters above bars indicate per-site significance groupings as inferred from Tukey HSD analyses.
12
sites (pANOVA, p N 0.05 for all sites) except for the site of Mnemba
where a constant decline in coral cover had been recorded (pANOVA,
p b 0.01), displaying the sharpest reduction of coral cover between
March 2007 and April 2008. Of the western sites, Chumbe recorded
the highest coral cover (64%–72%) throughout the three years of the
survey. The lowest coral cover was recorded on Kiwengwa, with cover
less than 0.5%. Macroalgal cover presented an opposite trend to coral
cover (Fig. 2B). While the western sites presented negligible macroalgal
cover, the eastern sites presented significantly higher algal cover
(pANOVA, p b 0.001), reaching more than 50% at some locations
(Fig. 2B). The highest macroalgal cover throughout the duration of the
surveys (ca. 35%–52%) was recorded on Kiwengwa on the Island's eastern exposed side. Temporal variations in macroalgal cover indicate no
change at the western sites (pANOVA, p N 0.05 for all sites), and significant increases in percentage cover for both Mnemba and Kiwengwa
(pANOVA, p b 0.001 for all; Fig. 2B). In the Mnemba MPA macroalgal
cover increased seven-fold from March 2007 to March 2008, and at
Pongwe in macroalgal cover constantly increased from ca. 1% to 9%
and 33% from 2006 to 2008.
Analyses of size frequency distributions revealed significant differences in size distributions (Table S1 in Electronic Supplementary
Material; ESM) and average sea urchin sizes among species (Table 1)
and sites (Table 2, Fig. S1). Sea urchins from the eastern sites were larger
than their western conspecifics. For example, the mean diameter of
E. diadema was highest on Pongwe, followed by Changu and Bawe
(ca. 100, 96, and 78 mm, respectively) (pANOVA, p b 0.001). Similarly,
E. mathaei presented relatively small individuals on Changu and Bawe
(ca. 36 and 34 mm, respectively), while those from Kiwengwa and
Pongwe were significantly larger (47 and 38 mm, respectively)
(pANOVA, p b 0.001). D. savignyi was significantly larger on Bawe
compared to Changu and Pongwe (pANOVA, p b 0.001). D. setosum presented significantly smaller individuals than their conspecifics on
Changu and Bawe (pANOVA, p b 0.002). Sea urchin densities on the
eastern MPA at Mnemba were too low to adequately perform size frequency estimations.
In all four species, the inorganic portion was larger than the organic
portion (Fig. 4). Interspecific differences were found in the CaCO3 and
organic matter portions, where E. diadema had a significantly larger proportion of organic matter and a smaller proportion of CaCO3 than all
other species (pANOVA, p b 0.01). D. setosum presented significantly
higher CaCO3 content and lower organic matter than all other species
(pANOVA, p b 0.001). No significant differences in the non-soluble portions were observed among species (pANOVA, p N 0.05).
Intraspecific comparisons of gut contents revealed significant differences in the proportions of CaCO3 and organic matter among species between sites (Fig. 4). However, while species like E. mathaei presented
increased proportions of organic matter on exposed reefs in comparison
to sheltered reefs, the results for the other species were not so clear. The
proportion of organic gut content of E. diadema on Changu was about
40%, compared to less than 20% on adjacent Bawe. The organic content
in the gut of E. diadema on the only exposed reef with sufficient sea urchin abundances, Pongwe, was around 20%. D. savignyi displayed a reverse pattern with Bawe around 25% organic gut composition
compared to 10% on Changu and 15% on Pongwe. D. setosum presented
similar CaCO3 and organic matter proportions on all sheltered sites
(~10%) (Fig. 4).
3.3. Echinoid gut contents analysis
Sea urchin populations varied in species assemblages and population densities both within and between sites, and across the western
and eastern sides of the Island (Fig. 3). Variations in densities were significant between sites (pANOVA, p b 0.001), and species (pANOVA,
p b 0.001). D. setosum dominated the western sheltered sites (i.e.
Changu, Bawe and Chumbe) followed by E. mathaei which instead dominated the exposed eastern ones (i.e. Kiwengwa and Pongwe) (Fig. 3).
E. mathaei densities were more than seven-fold higher on the eastern
sites than on the western ones (ca. 14 and 2 ind m− 2, respectively),
but were always absent from marine-protected areas on both sides.
The two MPAs, Mnemba (east), and Chumbe (west), differed from
neighboring sites in both species assemblages and densities (Fig. 3).
For the eastern reefs Mnemba presented low sea urchin densities
(1.62 ± 1.0 ind m− 2) and no E. mathaei, a sharp contrast to the
high echinoid densities at Kiwengwa and Pongwe (20.50 ± 12.0 and
30.19 ± 10.6 ind m− 2, respectively) (pANOVA, p b 0.001; Fig. 2).
On the western side, Chumbe had low density of E. mathaei compared
to Changu and Bawe (pANOVA, p b 0.05) but similar densities of
D. setosum (pANOVA, p N 0.05).
# of individuals m-2 (AVG ± SE)
40
E
W
A
35
a
AB
30
a
25
20
BC
15
ab
CD
CD
10
a
a
D
b
5
b
b
0
a
a
b b
Changu
b
b b
Bawe
b b b b
b b
Chumbe
b b
Mnemba
b b
b b
Kiwengwa
b
b
b
Pongwe
Sites
Diadema setosum
Diadema savignyi
Echinometra mathaei
Echinothrix diadema
Other
Fig. 3. Sea urchin densities at the different study sites around the Island of Zanzibar. Densities were measured in 20 m × 0.5 m belt transects (n = 10 transects per site). Bars indicate average species specific densities per m−2 (mean ± SD) from surveys conducted in March 2008. Western sites and eastern sites are denoted W and E, respectively. Significance groupings as
inferred from Tukey HSD analyses are presented as lowercase letters to indicate groupings among species within sites, and as uppercase letters among sites.
13
Table 1
Test diameters, gut dry weights, daily rates of food ingestion, herbivory and CaCO3 ingestion rates, and bioerosion rates for the mean-sized echinoid (mean ± SE) species around Zanzibar.
Diadema setosum (n = 30), Diadema savignyi (n = 30), Echinothrix diadema (n = 30) and Echinometra mathaei (n = 40). *Data obtained from Carreiro-Silva and McClanahan, 2001.
Species
Test diameter Gut weight
(mm)
(g)
Gut turnover
rate (day−1)*
Percentage of
reworked sediment
in the urchins gut*
Total ingestion
rate (gut dry weight
ind−1 day−1)
Herbivory rate
(g algae ind−1 day−1)
CaCO3 ingestion rate
(g CaCO3 ind−1 day−1)
Bioerosion rate
(g CaCO3 ind−1 day−1)
D. setosum
D. savignyi
E. diadema
E. mathaei
45.27
53.76
91.55
38.81
1.18
0.89
1.14
1.72
73.5
72.0
69.7
0
4.03
2.53
15.42
0.62
0.40
0.50
3.87
0.13
3.59
2.00
11.43
0.48
0.95
0.56
3.46
0.48
±
±
±
±
0.34
0.32
1.97
0.04
3.42
2.84
13.53
0.36
±
±
±
±
2.20
1.10
2.30
1.65
±
±
±
±
3.4. Rates of ingestion, herbivory, and bioerosion
Significant differences in rates of ingestion, herbivory and bioerosion
were found between sites and species. The greatest rates of ingestion
were found for E. diadema, with a daily CaCO3 consumption rate
of (11.43 ± 1.6 g CaCO3 ind− 1 day− 1), more than three-fold that
of D. setosum (3.59 ± 0.4 g CaCO3 ind−1 day−1), 5.7-fold that of
D. savignyi (2.00 ± 0.2 g CaCO3 ind−1 day−1), and almost 24-fold that
of E. mathaei (0.48 ± 0.05 g CaCO3 ind−1 day−1) (Table 1). E. diadema
also had the greatest rates of herbivory per day (3.87 ± 0.8 g algae
ind− 1 day−1) than any of the other species; 7.7-fold more than
D. savignyi (0.50 ± 0.1 g algae ind− 1 day− 1), 9.7-fold more than
D. setosum (0.40 ± 0.07 g algae ind−1 day−1), and almost 30-fold
more than E. mathaei (0.13 ± 0.02 g algae ind− 1 day− 1). E. diadema
also presented the highest individual rate of bioerosion (3.46 ± 0.5 g
CaCO3 ind−1 day−1). D. setosum average bioerosion rate (0.95 ± 0.1 g
CaCO3 ind−1 day−1) was less than a third of that of E. diadema while
almost twice the rate of D. savignyi and E. mathaei (0.56 ± 0.05 and
0.48 ± 0.05 g CaCO3 ind−1 day−1, respectively) (Table 1).
Gross annual bioerosion was higher on the eastern exposed reefs
in comparison to the western sheltered ones (ANOVA, F = 52.97,
p b 0.001; Fig. 5). The highest bioerosion rates were recorded on Pongwe,
with more than 6.9 kg CaCO3 m−2 eroded annually. Kiwengwa presented
the second highest bioerosion levels (4.2 kg CaCO3 m−2 year−1), slightly
higher than Bawe (3.9 kg CaCO3 m−2 year−1) and twice the erosion rates
at Changu (2.1 kg CaCO3 m−2 year−1). The lowest bioerosion rates were
recorded on Chumbe, with only 1.2 kg m−2 year−1 CaCO3 eroded
(Table 2, Fig. 5).
4. Discussion
The magnitude of echinoid grazing, and consequently the rates of
herbivory and bioerosion, are believed to be determined by three
0.41
0.29
2.25
0.06
±
±
±
±
0.07
0.13
0.78
0.02
±
±
±
±
0.37
0.18
1.59
0.05
±
±
±
±
0.10
0.05
0.48
0.05
fundamental variables: species identity, body size and population densities (Bak, 1994). These variables are both governed by the habitat's physical conditions and at the same time, influence the structure and
composition of the habitat itself (reviewed by Steneck, 2013). For example, factors such as habitat structural complexity (Graham and Nash,
2013), exposure to surf (Ebert, 1982), and regulation by predation
(McClanahan and Kurtis, 1991; McClanahan and Shafir, 1990), are all
known to influence echinoid community structure on coral reefs. On
the other hand, reef degradation through bioerosion is often associated
with high sea urchin abundance (Bak, 1990; Scoffin et al., 1980), direct
coral predation (Carpenter, 1981; Glynn et al., 1979), and echinoid
control of newly-settled spat (Sammarco, 1980, 1982). The type and intensity of environmental conditions are likely to regulate echinoid communities both directly (e.g., inability of certain species to resist strong
currents) (Tuya et al., 2007), and indirectly through shaping coral reefs
structure to either favor or exclude specific echinoid species (Dumas
et al., 2007; McClanahan and Kurtis, 1991; McClanahan and Shafir,
1990).
4.1. Coral and macroalgal cover
We found marked differences in coral and macroalgal cover, echinoid community structure, and rates of herbivory and bioerosion between exposed and sheltered reefs. That coral cover is low in areas of
high algal cover has been thoroughly discussed in the scientific literature, and is often associated with algal predominance over scleractinian
corals in competition for environmental resources (reviewed by
McCook et al., 2001). Respectively, our data present opposite trends of
coral and algal cover: where coral cover was low, algal cover was high
and vise versa. This pattern was evident in all of the sites studied regardless of the level of ocean exposure or protection (i.e. MPA and nonMPA), reflecting the generality of these coral–algal interactions. Nonetheless, between-site differences corresponded to the degree of habitat
Table 2
Densities, test diameters, herbivory rates, CaCO3 ingestion rates, and bioerosion rates (mean ± SE) of the dominant echinoid species of Zanzibar. Data presented by site and are based on
the average test diameters of each species in every site (n = 10 individuals per species in each site).
Site
Sea facing/
category
Species
Density
(ind m−2)
Test diameter
(mm)
Herbivory rate
(g algae ind−1
day−1)
CaCO3 ingestion
rate (g CaCO3
ind−1 day1)
Bioerosion rate
(g CaCO3
ind−1 day−1)
Net herbivory
(kg algae
m−2 year−1)
Net bioerosion
(kg CaCO3
m−2 year−1)
Total site
bioerosion
(kg CaCO3
m−2 year−1)
Changu
West (sheltered)
West (sheltered)
Chumbe
Kiwengwa
Pongwe
West (sheltered)
East (exposed)
East (exposed)
Mnemba
East (exposed)
5.99
0.44
0.24
1.26
5.72
0.4
0.18
4.79
6.38
20.28
1.65
1.41
12.47
NA
46.45
52.69
96.19
35.65
47.52
55.55
78.31
34.03
41.85
47.42
53.06
100.16
38.14
NA
0.31
0.15
2.47
0.05
0.69
1.07
1.62
0.04
0.20
0.24
0.29
7.52
0.17
NA
2.97
1.36
3.75
0.46
5.82
2.55
8.38
0.23
1.97
0.57
2.08
22.17
0.68
NA
0.79
0.38
1.14
0.46
1.54
0.72
2.54
0.23
0.52
0.57
0.58
6.72
0.68
NA
0.68
0.02
0.22
0.02
1.44
0.16
0.11
0.07
0.47
1.78
0.17
3.87
0.77
NA
1.73
0.06
0.10
0.21
3.22
0.11
0.17
0.40
1.21
4.22
0.35
3.46
3.10
NA
2.10 ± 0.08
Bawe
D. setosum
D. savignyi
E. diadema
E. mathaei
D. setosum
D. savignyi
E. diadema
E. mathaei
D. setosum
E. mathaei
D. savignyi
E. diadema
E. mathaei
NA
±
±
±
±
±
±
±
±
±
±
±
±
±
0.56
0.07
0.01
0.36
0.87
0.06
0.12
2.80
0.57
3.63
0.57
0.26
3.98
±
±
±
±
±
±
±
±
±
±
±
±
±
1.00
0.53
0.85
0.80
1.15
0.55
0.60
0.84
1.36
0.86
0.68
0.99
1.10
±
±
±
±
±
±
±
±
±
±
±
±
±
0.04
0.01
0.72
0.01
0.17
0.33
0.55
0.00
0.02
0.02
0.02
1.69
0.02
±
±
±
±
±
±
±
±
±
±
±
±
±
0.26
0.11
1.09
0.10
0.53
0.39
0.63
0.04
0.26
0.10
0.30
1.56
0.12
±
±
±
±
±
±
±
±
±
±
±
±
±
0.07
0.03
0.33
0.10
0.14
0.11
0.19
0.04
0.07
0.10
0.08
0.47
0.12
±
±
±
±
±
±
±
±
±
±
±
±
±
0.09
0.00
0.06
0.00
0.36
0.05
0.03
0.00
0.04
0.14
0.02
1.69
0.16
±
±
±
±
±
±
±
±
±
±
±
±
±
0.15
0.00
0.03
0.05
0.29
0.02
0.01
0.07
0.16
0.23
0.09
0.47
0.99
3.90 ± 0.15
1.21 ± 0.16
4.22 ± 0.23
6.91 ± 0.63
NA
14
Changu
80
60
40
20
0
100
Bawe
80
60
40
20
0
100
NA
NA
NA
40
20
0
E
W
6
4
2
0
Changu
Bawe
Chumbe Mnemba Kiwengwa Pongwe
100
80
NA
60
NA
NA
40
20
Kiwengwa
% of gut contents
60
Chumbe
80
Bioerosion rates (kg CaCO3m−2 yr−1) (AVG ± SE)
100
0
80
NA
60
40
Pongwe
100
20
0
100
All sites
80
60
40
20
0
E.mathaei
E.diadema
D. savignyi
D. setosum
Species
Calcium carbonate
Non - soluble
Sites
Diadema setosum
Diadema savignyi
Echinometra mathaei
Echinothrix diadema
Fig. 5. Echinoid annual bioerosion rates per site and species at six sites surrounding the
Island of Zanzibar. Bars indicate mean ± SE bioerosion rates per m−2 of each species
(color coded) and every site (stacked). Western sites and eastern sites are denoted
W and E, respectively. The number of samples collected was n = 10 per species per site.
of herbivorous fish may compensate for the lack of echinoids in regulating algal cover at that site. Thus, though the reason for Mnemba's coral
cover loss is not resolved at this stage, it might be a consequence of global changes and the overall large-scale coral decline (Bellwood et al.,
2004; Hughes, 2003). In contrast coral cover on the Chumbe MPA was
exceptionally high throughout the duration of the study. Some of the
mechanisms that work to maintain such high coral cover may be attributed to the extensive and efficient protection these reefs receive (e.g.,
no take zone, diving restrictions and strict anchoring regulations). However, in contrast to the reefs on Changu and Bawe, the geographical location of Chumbe, away from the big urban center of Zanzibar Town, is
likely to further reduce indirect anthropogenic stressors, such as water
pollution, that may also contribute to the prosperity of the latter reefs.
Organic matter
Fig. 4. Fractionation of gut contents (mean ± SE) at each study site, and the pooled data
per species of the four dominant echinoid species on the coral reefs of Zanzibar. Color indicates: CaCO3 (dark gray), non-soluble residues (light gray) and organic matter (black)
fractions. (n = 10 individuals per species, per site).
exposure: while exposed reefs had high algal and low coral cover,
sheltered reefs presented an opposite trend. It appears that coral communities on Zanzibar's exposed reefs are being competitively excluded
by algae, while on sheltered reefs coral communities are sustained
through algal regulation. Of particular interest is the Mnemba MPA on
the north-eastern exposed side. Our data show a significant decrease
in coral cover and, at the same time, a significant increase in macroalgal
cover between 2007 and 2008 (Fig. 2). Though coral cover at this site is
historically low (Bergman and Öhman, 2001; Ngoile, 1990), our measurements reflect the lowest ever recorded coral cover in what is considered Zanzibar's oldest established MPA (Obura et al., 2002).
Nonetheless, management at this site seems effective, judging by the
abundance of herbivorous and predatory fish on the site (Brokovich,
pers. comm.). The abundance of predatory fish may account for the
lack of sea urchins (McClanahan and Kurtis, 1991), while the abundance
Multiple environmental, ecological, and biological factors concurrently occurring on coral reefs make it difficult to elucidate the resulting
echinoid species distributions (Dumas et al., 2007; Johansson et al.,
2013). Nonetheless, echinoid community structure in terms of species
composition, densities, and average body size, varied considerably between sheltered and exposed reefs. Of the four dominant echinoid species found around the Island of Zanzibar, the two most abundant
species, D. setosum, and E. mathaei, were also the most affected by the
level of habitat exposure. While D. setosum was prevalent on sheltered
sites, it was absent on exposed sites where E. mathaei predominated
(Fig. 3). D. setosum is known to be restricted to quiet or protected waters
(Ebert, 1982). In exposed areas, the large overall volume (body and
spines) of Diadema may prevent it from resisting the strong currents
and surf, causing it to detach from the substrate, as demonstrated in
Diadema on the reefs of the Canarian Archipelago (Tuya et al., 2007). Despite high macroalgal cover (Fig. 2B) and presumably low predation
pressure as a result of intense fishing activity in the area (Jiddawi and
Öhman, 2002; Ngoile et al., 1988), D. setosum spatial distribution is
most likely a result of physical variables (e.g., surf, currents, etc.) rather
than biotic factors (e.g., predation, food availability, etc.).
In contrast to D. setosum, E. mathaei, though present at most western
sites, predominated the eastern exposed sites (Fig. 3). Similar to Russo's
findings from reefs with stronger water flow (Russo, 1977), our data
show higher densities and a larger average body size of E. mathaei on
the exposed reefs. These patterns of size and abundance might be attributed to food limitations and availability in the two different reef
environments. As E. mathaei is generally sedentary, limited to the proximity of its burrows (Langdon, 2012; Young and Bellwood, 2011), in
late successional algal communities it was suggested to rely on drift
algae for a large part of its diet (Johansson et al., 2013; McClanahan and
Muthiga, 2001; Russo, 1977). The significantly higher macroalgal cover
and stronger currents on the eastern exposed reefs generate large
amounts of drift algae, thus increasing food accessibility for populations
of E. mathaei. However, in comparison to D. setosum, E. mathaei's distribution is thought to be less susceptible to environmental variables (Dumas
et al., 2007). E. mathaei's proliferation throughout the exposed unprotected sites might also be attributed to its ability to competitively exclude
other echinoid species (McClanahan, 1988a). In this context of interspecific competition, variations in the echinoid guild of the Bawe site
will be interesting to follow during the next few years, as E. mathaei densities there are constantly rising (Bronstein, unpublished data) and are
now as high as those of D. setosum (Fig. 3, Bronstein, unpubl. data).
Regardless of the level of exposure, no E. mathaei were recorded on
the two MPAs, Chumbe and Mnemba. Predation is probably the prevalent regulatory agent affecting E. mathaei proliferation on marineprotected areas (McClanahan and Kurtis, 1991; McClanahan and
Muthiga, 1989). Exclusion by predation can thus explain the sharp contrast between the extremely high E. mathaei densities observed on reefs
adjacent to the Mnemba MPA, in contrast to their absence from within
the protected zone. Alternatively, the effects of predation might still
be evident even when echinoids are not completely excluded as in the
case of Mnemba. Strong predation pressure of D. setosum in the MPA
of Chumbe could be assumed from their bimodal size frequency distribution (Fig. S1), as mid-sized individuals are most susceptible to predation while large predator-immune sea urchins and newly recruited
individuals are more likely to escape predation (Ojeda and Dearborn,
1991; Scheibling and Hamm, 1991; Tegner and Levin, 1983). In unprotected reefs, the difference in E. mathaei proliferation between exposed
and sheltered reefs may also be attributed to human exploitation
through fishing. Reduced fishing success on sheltered reefs may be
attributed to the relatively high coral cover and structural complexity
of these reefs, which may provide more refuge from fishing for potential
E. mathaei predators in comparison to bare exposed reefs (McClanahan,
1997). Alternatively, the strong surf and currents associated with exposed reefs may restrict fishing by forcing fishermen away from exposed sites or limit fishing duration, allowing more echinoid predators
to avoid being caught.
4.3. Bioerosion and herbivory
Accurate assessments of the proportion of reworked sediments in
the diet of echinoids are still inconclusive, as evident from the variety
of methodologies that have been used to obtain them. These methodologies, often yielding considerably different results, may include the use
of petrographic sections of fecal pellets (Bak, 1990; Hunter, 1977;
Scoffin et al., 1980), CaCO3 evaluations in the guts of urchins from
noncarbonated substrates (Mokady et al., 1996), and even comparisons
to other echinoids that are presumed to be non-eroding species
(Carreiro-Silva and McClanahan, 2001). Nonetheless, except for the
biases that originate from using different methodologies, it is still largely
debated whether ingestion of loose sediment (reworked sediment) by
sea urchins is merely a by-product of their grazing activity (i.e. unintentional sediment ingestion) and as such a consequence of sediment loads,
or an active strategy of intentional grazing that is governed by the life
histories of the different species. In the current study, we have observed,
on several occasions, groups of Diadema setosum actively feeding on
15
filamentous algae on loose sediments (Bronstein, unpublished data).
Such activity would imply a nutritional strategy for utilizing an available
resource through intentional sediment ingestion. These observations,
despite being far from conclusive, should serve to underline gaps in
our understanding of this process, and stress the need to further elucidate the issue of echinoid reworked sediment evaluations.
Sheltered sites experienced lower degrees of bioerosion in comparison to exposed sites. Of the sheltered sites, the highest degree of
bioerosion was recorded on Bawe, being over three-fold higher than
that recorded on Chumbe, and almost twice as that of Changu
(Table 2). Some of these differences may be attributed to differences
in species composition and abundance on these reefs. While in Bawe
four different echinoid species contributed to the total bioerosion, in
Chumbe D. setosum was the sole contributor, and although similar species compositions were observed on Bawe and Changu, the abundance
of E. mathaei on Bawe was 3.8 times higher than on Changu. Gross
bioerosion on Bawe is still more than two-fold higher than that on
Chumbe, even when only D. setosum is considered, and despite similar
abundance and algae cover on Chumbe. Apparently the higher gross
bioerosion on Bawe may instead be attributed to the higher average
body size of D. setosum on these reefs, as rates of food consumption
are known to increase with sea urchin size (Bak, 1990; Klumpp et al.,
1993). However, this argument seems inadequate here as D. setosum
on Bawe present rates of bioerosion almost twice as high as those on
Changu, despite having similar average body size. Other currently unaccounted for variables, such as the hardness or particular type of substrata available for grazing, may account for these differences.
Bioerosion rates on the exposed sites were highest on Pongwe, with
more than 6.9 kg CaCO3 m−2 eroded annually. The higher gross
bioerosion on Pongwe compared to Kiwengwa, despite the higher total
sea urchin abundance on the latter site (ca. 15.5 and 20.3 ind m−2, respectively), may be attributed to differences in species composition between the two sites. While on Pongwe there are three species
comprising the reef-eroding echinoid guild, the reefs of Kiwengwa are
solely occupied by E. mathaei. As different species may occupy different
niches on the reef, interspecific competition may be reduced allowing
more resources for feeding. Although intraspecific food consumption is
expected to increase as sea urchins grow larger (Bak, 1990; Klumpp
et al., 1993), the individual rates of food consumption for the larger
mean sized E. mathaei from Kiwengwa were lower than those from
Pongwe. The high E. mathaei abundance on Kiwengwa inevitably increases intraspecific competition which consequently utilizes resources
that might otherwise be channeled to feeding.
The main differences in bioerosion between the eastern and western
reefs should rather be attributed to E. mathaei, which total erosion was
more than 12 times higher on eastern reefs. These rates are higher
than those reported for E. mathaei from the Red Sea 0.5–0.9 kg CaCO3
m−2 year−2 (Mokady et al., 1996); Enewetak Atoll 0.08–0.33 kg
CaCO3 m−2 year−2 (Russo, 1980); Marshall Islands 3.3 kg CaCO3 m
−2
year−2 (Russo, 1980); Kenya 1.3 kg CaCO3 m−2 year−2 on unprotected reefs (Carreiro-Silva and McClanahan, 2001); and Moorea, French
Polynesia 0.37 kg CaCO3 m−2 year−2 (Bak, 1990), but lower than the
exceptionally high rates reported from the Arabian Gulf 9.9–15.3 kg
CaCO3 m−2 year−2 (Downing and El-Zahr, 1987). Another difference
between exposed and sheltered reefs is the absence of D. setosum
on the eastern reefs. However, the lack of D. setosum bioerosion is
diminished by the high bioerosion rates of E. mathaei.
One important contributing factor to different bioerosion rates is
body size. The larger body size of E. diadema may act in two ways to facilitate its exceptionally high herbivory and CaCO3 consumption rates
(Table 2): (a) the larger Aristotle's lantern (the sea urchin's feeding apparatus) and intestines, associated with larger body size (Black et al.,
1982; Ebert, 1980), may increase the volume of food ingestion and digestion; and (b) a larger body may reduce the risk of predation
(McClanahan and Muthiga, 1989), increasing the duration of food foraging at the expense of seeking shelter. In addition to quantity, body size
16
may also determine the quality of the food consumed. For example, in
two similarly-sized Diadema populations (67.8 ± 6.2 mm and 69.9 ±
6.1 mm, mean ± SD, for D. setosum and D. savignyi, respectively) in
Kenya, no difference was observed in the organic and CaCO3 portions
of the two species (Carreiro-Silva and McClanahan, 2001). In contrast,
a significant size difference in similar species in Zanzibar (45.3 mm
and 53.8 ± 0.3 mm, mean ± SE, for D. setosum and D. savignyi, respectively), revealed a higher proportion of organic matter in the larger
D. savignyi. As larger Diadema individuals are potentially more mobile
and less prone to predation, they may exploit feeding grounds inaccessible to smaller individuals, consequently maximizing organic matter
consumption.
The two MPAs, Chumbe on the west and Mnemba on the east, differed from neighboring unprotected sites in both echinoid species composition and abundance, and consequently in the rates of bioerosion.
For example, on Mnemba, echinoid densities were so low that seaurchin-induced bioerosion was considered negligible. Similarly on
Chumbe, despite the presence of D. setosum, bioerosion rates were the
lowest recorded per site (excluding Mnemba). As predation controls
echinoid proliferation (McClanahan, 1998), and fishing may in turn
reduce potential predators (McClanahan et al., 1999), sea urchin populations on highly exploited reefs are expected to proliferate. Consequently, except for protected areas, where fishing restrictions are
enforced, extensive reef exploitation may have favored sea urchin proliferation, which is accelerating reef degradation through bioerosion.
In addition to regulation of community structure, the type of substratum has also been shown to directly affect echinoid rates of
bioerosion. Hibino and Van Woesik (2000), for example, found substrate age to be an important factor affecting bioerosion rates by
E. mathaei: older substrates lost on average more carbonate than did recently dead coral on experimental tiles. As echinoid populations on
sheltered reefs, in the current study, inhabit fairly flourishing reefs,
they encounter mostly young substrata (mostly coral and coralline
algae) of recently accreted origin. In contrast, the exposed reefs on
Zanzibar's eastern side are showing continued signs of deterioration
and loss of hard-coral cover (Mbije et al., 2002). Thus, with little accretion to compensate for the loss of substrata, echinoid bioerosion is expected to accelerate with time, as older and faster eroding parts of the
reef are constantly being exposed. Apart from sea urchins already
being key drivers of the carbonate budget on most coral reefs (Scoffin
et al., 1980), as coral cover declines the system will become less sensitive to drivers of calcification and ever more sensitive to echinoid
bioerosion (Perry et al., 2012). Moreover, intense erosion of CCA will
not just eliminate a significant ingredient of the reef framework, but
may also hinder reef recovery as some species are known to chemically
induce recruitment of corals and octocorals (Harrington et al., 2004;
Heyward and Negri, 1999; Morse et al., 1996).
Sea urchin erosion may temporarily act to increase the complexity of
exposed reefs by enhancing to the 3-dimensional structure of abraded
benthic environments, consequently facilitating increased diversity in
these habitats. However, as echinoid populations continue to proliferate, resources will eventually dwindle, and intensifying competition
might work to exclude other benthic species, leading to reduced biodiversity. Consequently, in areas of low coral cover and little regulation of
echinoid proliferation, such as on Zanzibar's exposed unprotected reefs,
sea urchin bioerosion may ultimately lead to a negative carbonate budget and consequently to reef degradation. Our data suggest that these
processes are happening faster on exposed rather than sheltered reefs.
These differences may be attributed to the complete dominance of the
sea urchin E. mathaei, having higher densities and a larger average
body size than on sheltered reefs. Nonetheless, the establishment of marine protected areas can dramatically slow this process, and facilitate
the sustainability of the reef framework by providing efficient regulation over echinoid populations.
Supplementary data to this article can be found online at http://dx.
doi.org/10.1016/j.jembe.2014.03.003.
Acknowledgments
This work was supported by the World Bank Group and the Global
Environmental Facility (GEF) through the Coral Reef Targeted Research
and Capacity Building for Management program, Coral Bleaching
and Local Environmental Responses working group to YL. We thank
the staff of the IMS in Tanzania for the use of their facilities. We are
grateful to Dr. A. Zvuloni for providing data on coral and macroalgae,
to Dr. N. Shenkar for her advice on an early draft of this manuscript,
and to N. Paz and M. Chen Bronstein for their editorial assistance. [ST]
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Echinoid community structure and rates of herbivory and bioerosion

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