Marine Pollution Bulletin 62 (2011) 1140–1146
Contents lists available at ScienceDirect
Marine Pollution Bulletin
journal homepage: www.elsevier.com/locate/marpolbul
Baseline
Sponges and sediments as monitoring tools of metal contamination
in the eastern coast of the Red Sea, Saudi Arabia
Ke Pan, On On Lee, Pei-Yuan Qian, Wen-Xiong Wang ⇑
Division of Life Science, The Hong Kong University of Science and Technology (HKUST), Clear Water Bay, Kowloon, Hong Kong
a r t i c l e
i n f o
Keywords:
Sediments
Sponges
Red Sea
Metals
Saudi Arabia
a b s t r a c t
Sediments and sponges were collected from various locations along the eastern coast of the Red Sea, the
Kingdom of Saudi Arabia. Total concentrations of Cd, Zn, Ag, Cu, Pb, As and Hg in the sediments were measured. Metal contamination was not significant in most of the studied sites and only one site was moderately polluted by Zn, Cu, and Pb. Sponges accumulated specific metals readily even though the metal
exposure was low in the ambient environment. Contrasting interspecies differences in metal accumulation patterns were observed among the nine collected species of sponges. Significant positive correlations
were found between the metal concentrations in the two species of sponges collected from the same
sites. The strong ability to accumulate specific metals and the diversity of sponges that live in the Red
Sea coastal areas make them a promising biomonitor of metal contamination in the areas.
Ó 2011 Elsevier Ltd. All rights reserved.
Coastal areas are often considered as receptacles for pollutants
from industrial and urban activities. Overexploitation in modern
society has led to elevated inputs of anthropogenic contaminants
into coastal and estuarine areas which are vulnerable to human
activities (Peters et al., 1997; Lewis and Devereux, 2009). Metals
are one of the toxic pollutants of great concern because of their potential detrimental effects on aquatic ecosystems (Grimalt et al.,
2001; Prouty et al., 2010). Meanwhile, metals are readily accumulated by aquatic organisms and subsequently transferred along
food chains, ultimately posing a risk to human health through seafood consumption (Wang, 2002; Gerstenberger et al., 2010).
The Red Sea is famous for its unique tropical coral reef, mangrove, and seagrass ecosystems (Odum and Odum, 1955; Price
et al., 1998; Ashworth et al., 2006). The coastline of the Kingdom
of Saudi Arabia stretches for about 1840 km and accounts for 79%
of the eastern coast of the Red Sea. This area provides various habitats for diverse communities of corals and sponges. As one of the
largest countries bordering the Red Sea, Saudi Arabia has undergone a rapid transformation into a modern industrial country (Badr
et al., 2009). As a result, a significant part of the coast has been subjected to extensive exploitation, and metal pollution is fast becoming a threat to the coastal environments. Incidents of damaged oil
wells, oil pipeline leaks, and domestic sewage from coastal cities
are contributing significantly to the coastal pollution (Al-Thukair
et al., 2007). Kadi (2009) showed that soils in the urban areas of
Jeddah—a Saudi Arabian city located on the coast of the Red
Sea—have been polluted by Zn and Pb found in traffic road dust.
⇑ Corresponding author. Tel.: +852 23587346; fax: +852 23581559.
E-mail address: [email protected] (W.-X. Wang).
0025-326X/$ - see front matter Ó 2011 Elsevier Ltd. All rights reserved.
doi:10.1016/j.marpolbul.2011.02.043
The highest concentration found in roadside soils was 105 mg kg1
for Pb and up to 450 mg kg1 for Zn. Recent records from sediment
cores also revealed that Jeddah was the most polluted area along
the eastern coast, where over 3 mg kg1 of Cd and 100 mg kg1
of Pb was detected in the sediments and increasing metal concentrations were observed in the upper layer of the cores (Badr et al.,
2009). To date, limited data exist for an accurate assessment of the
metal pollution of coastal environments in Saudi Arabia, especially
for the areas located near the coral reef and mangrove ecosystems.
Field surveys of metal pollution must be conducted before appropriate policies can be made to protect the vulnerable coastal
environments.
Investigating the concentration and distribution of metals in
sediments is an effective way to understand metal contamination
in marine ecosystems because sediments are a reservoir for metals
and can provide historical input records of metals (Hatji et al.,
2002; Bell et al., 1997). Sediments are preferred as a monitoring
tool because they generally show less variation over a short period
of time than dissolved metals in overlying water columns
(Atkinson et al., 2007). Although the total metal concentration is
a valuable piece of information and tells us about the integrated
accumulation of metals in sediments over a certain period of time,
it is however inadequate to predict the mobility, bioavailability
and potential toxicity of metals in hazard assessment. The fates
of metals in sediments are greatly dependent on their physicochemical speciation and environmental conditions such as redox,
pH, salinity, and temperature (Yu et al., 2010). Sediment geochemistry can significantly control metal bioavailability. Previous studies have shown that labile metals (such as metals bound in
exchangeable phase) are more bioavailable than those bound with
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K. Pan et al. / Marine Pollution Bulletin 62 (2011) 1140–1146
the sulfide species or the reducible phase (Chen and Mayer, 1999;
Stecko and Bendell-Young, 2000; Fan et al., 2002a). Therefore,
measurement of geochemical species of metals is a necessary supplement to the environmental assessment of contaminated
sediments.
Knowing the characteristics of either seawater or sediments is
not enough to predict unequivocally the bioavailability of metals.
Benthic species such as bivalves, seagrasses, macroalgae and
sponges have been utilized as biomonitors to indicate metal availability in marine environments (Roberts et al., 2008). Sponges—
animals of the phylum Porifera—have been recommended as a
suitable biomonitor species for metal pollution because of their
strong ability to concentrate metals in their tissues (Berthet
et al., 2005; Patel et al., 1985; Johnston and Clark, 2007; Cebrian
et al., 2007). They are one of the major benthic groups with a prominent role in many coral reef communities around the world,
including in the Red Sea ecosystems (Ilan et al., 2004). Sponges
can not only absorb dissolved metals but also take up particulate
metals by filtrating suspended matters. Meanwhile, sponges themselves act as biogenic habitats that support abundant and highly
diverse epifaunal and infaunal microbial communities which make
up significant biomass of their host. To date, the number of studies
employing sponges as biomonitors for metal contamination in the
Red Sea is still limited despite their huge potential.
The aims of this study were therefore to perform a baseline
investigation of the state of metal pollution along the eastern coast
of the Red Sea by measuring the total metal concentrations and
geochemical speciation of metals in the sediments, and to evaluate
the use of sponges as biomonitors for metal contamination in the
area.
Sediments and sponges were collected from nine sites located
along the eastern coast of the Red Sea in Saudi Arabia in April
2009. Both pristine areas and areas affected by human activities
were included. The location of each site is shown in Fig. 1 and Table 1. The inner bay, marine station, treatment plant outfall, fish
market and its entrance area in Obhor Sharm were considered as
areas affected by human activities, whereas the bay outlet, the
mangrove site, Abu Madafi reef and the non-reef reference site in
Obhor Sharm were considered as pristine sites away from human
disturbance. Abu Madafi Reef is one of the most well-preserved
Fig. 1. Sampling sites along the eastern coast of the Red Sea, the Kingdom of Saudi
Arabia. Abbreviations are defined in Table 1.
Table 1
Location of the sampling sites along the coast of the Red Sea, the Kingdom of Saudi
Arabia.
Site
Site name
Abbreviation
Location
S1
S2
S3
S4
S5
S6
S7
S8
S9
Obhor Sharm (outlet)
Marine station
Obhor Sharm (inner)
Abu Madafi reef
Non-reef reference site
Mangrove site
Treatment plant outfall
Entrance of fish market
Fish market
OSO
MSN
OSI
AMR
NRR
MGR
TRO
EFM
FMT
210 42.323N
210 42.642N
210 45.687N
220 03.656N
220 10.178N
220 13.139N
210 19.400N
210 29.622N
210 29.260N
390 04.230E
390 05.685E
390 08.061E
380 46.074E
380 57.398E
390 03.069E
390 05.887E
390 09.617E
390 10.540E
coral reefs in the Red Sea; the non-reef reference site was selected
for comparisons. The top 3–5 cm of surface sediments and sponge
samples were collected by a scuba diver or a snorkeler at each site
(n = 3), depending on the water depth (1–30 m). Sediment samples
were stored at 4 °C, while sponge samples were stored at 20 °C
prior to metal analysis. Separate specimens of each sample were
preserved in 70% ethanol for species identification.
Upon transportation back to the laboratory, sediments were
freeze-dried and large stones and debris were removed. The fine
fraction of sediment (<63 lm) was separated by passing it through
a polyethylene sieve. The <63 lm fraction was used because this
fraction proved to be the most chemically active sediment phase,
consisting primarily of clay and silt particulates (Förstner, 1987).
The sponge samples were carefully rinsed with 0.22 lm filtered
seawater to clean off any foreign material such as loosely-bound
sediments and epibionts, and were later freeze-dried. Homogenized samples were made by cutting the sponges into small pieces
followed by grounding with a mortar and pestle.
To measure the total metal contents in the sediment samples,
approximately 0.2 g of the representative sample was placed into
a Teflon reactor, and then digested in a solution consisting of
2 mL of 70% Suprapur nitric acid (HNO3), 6 mL of 37% hydrochloride acid (HCl) and 100 lL of hydrofluoric acid (HF) at a temperature of 180 °C for 15 min in a microwave digestion system
(BERGHOFÒ Speedwave MWS-3, Germany). Approximately 0.2 g
of sponge samples were digested with 3 mL of 70% nitric acid
and 1 mL of H2O2 as described in previous studies (Cebrian et al.,
2007). The digested samples were measured for Cd, Zn, Ag, Cu,
As and Pb using an atomic absorption spectrometer (AAS, PerkinElmer, AAnalyst 800) or an inductively coupled plasma optical emission spectrometer (ICP-OES, PerkinElmer, Optima 7000 DV). Total
Hg concentrations were measured by employing QuickTrace™ M8000 Cold Vapor Atomic Fluorescence mercury analyzer (USA).
Cd, Zn, Cu, Ag, As and Hg were analyzed in this study because they
are common metal contaminants and reports of Ag and Hg values
for the Red Sea coasts are rare. The analytical accuracy was
checked by concurrent digestion and comparing measurements
with NIST reference materials: estuarine sediment (SRM 1646a)
and oyster (SRM 1566b). The recoveries were all within 90–110%
of the reference values and the data were not corrected for
recovery.
Metal speciation of Cd, Zn, Pb, Ag, and Cu was only measured in
the sand-silt texture samples collected from the inner bay, fish
market and its entrance area in Obhor Sharm (S3, S8 and S9). Metal
speciation in the sediments was quantified using the sequential
extraction method described in previous studies (Tessier et al.,
1979; Fan et al., 2002b). The five operationally defined geochemical fractions were as follows:
(1) Phase 1 (exchangeable fraction): extracted by placing in 1 M
MgCl2 at pH = 7 for 1 h.
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K. Pan et al. / Marine Pollution Bulletin 62 (2011) 1140–1146
(2) Phase 2 (carbonate bound fraction): extracted by placing in
1 M NaOAc at pH = 5 for 5 h.
(3) Phase 3 (Fe or Mn oxides bound fraction): extracted by placing in 0.04 M NH2OH-HCl in 25% HOAc for 6 h at 96 °C.
(4) Phase 4 (organic matter bound fraction): extracted by placing in 0.02 M HNO3 and 30% H2O2 (adjusted to pH = 2 with
HNO3) at 85 °C for 2 h, followed by the addition of 30%
H2O2 for another 3 h. After cooling, 3.2 M NH4OAc in 20%
HNO3 was added and the mixture was continuously agitated
for 30 min.
(5) Phase 5 (residual phase): extracted by placing in concentrated HNO3 (70%) and HClO4 (60%) and heated to dryness.
The remaining material was further digested in 10 ml of
2% HNO3 at 70 °C for 1 h.
Following each extraction in steps (1)–(4), the mixtures were
centrifuged at 4000 rpm for 30 min at room temperature and the
supernatants were heated to dryness in an aluminum heating
block at 140 °C. The remaining material was further digested in
10 mL of 2% HNO3 at 70 °C for 1 h. Recoveries of the metals were
checked by extracting the metals from NIST 1646a sediments using
the same procedure above. Metal concentrations of the extracted
solution were measured using the same AAS from PerkinElmer as
described above.
Statistical analysis was performed using the SPSSÒ 16.0 software package. All the data were tested for homogeneity of variance
and normal distribution before the statistical analysis. Statistically
significant differences among the sites/sponges were detected
through one-way analysis of variance using a least-significantdifference post hoc test (p < 0.05). Differences in metal concentration in sediments were assessed by one-way ANOVA. Differences in
metal accumulation between Hyrtios erectus and Stylissa carteri
were analyzed by two-way ANOVA. Regression analysis of the
metal concentrations in H. erectus and S. carteri was performed
using the software package SigmaPlot 9.0.
The total metal concentrations in the sediments collected from
nine sampling sites are summarized in Table 2. Huge differences
(up to 20 times) in metal concentrations were observed between
the least and the most contaminated site, showing the contrasting
variation in mineralogical composition or the different levels of
anthropogenic input at each site. The concentration range for each
metal was: 0.024–0.24 mg kg1 for Cd, 5.3–179.0 mg kg1 for Zn,
0.05–0.95 mg kg1 for Ag, 0.45–82.99 mg kg1 for Cu, 0.46–69.38
mg kg1 for Pb, 1.4–21.0 mg kg1 for As, and 3.0–132.8 lg kg1
for Hg. A relatively higher metal concentration was found in inner
Obhor Sharm bay (S3) and the adjacent areas of the fish market (S8
and S9), due possibly to the result of high shipping activities in the
areas. The fish market (S9) was the most polluted site out of all
nine sites sampled in terms of the total metal concentrations of
Cd, Zn, Ag, Cu, and Pb (p < 0.01). The pristine sites, the bay outlet
of Obhor Sharm, the mangrove site, Abu Madafi reef and the
non-reef reference site, accumulated much less metals in the sediments. The sediments deposited in the Red Sea coastal areas can
be classified into three principle categories: biogenous, terrigenous
and authigenic. Biogenous sediments are mainly composed of
eroded coral reefs and various calcareous remains from marine
organisms (Basaham, 2009). High carbonate contents are generally
associated with low concentrations of trace metals (Rubio et al.,
2000). The low metal contents found in S4 may be due to the higher proportions of biogenic carbonate and aragonite in the samples
that originated from the coral areas. When compared to numerical
sediment quality guidelines (SQGs, Long et al., 1998; MacDonald
et al., 2000), the concentrations of Zn, Cu, and Pb in the fish market
(S9) exceeded the threshold effect concentration (TEC, Table 2), but
barely reached or exceeded the median effect concentration (MEC).
The concentration of arsenic (As) in the mangrove site exceeded
the MEC, indicating possible As contamination in the area. The concentrations of metals in other sites were far below the TEC or the
average crust concentration. Generally, the metal concentrations
of most sites in our study were low, indicating that the local environment was less affected by metal contamination than other
industrialized coastal areas.
Metals were sequentially extracted from select sediment samples, all of which had a sand-silt texture, from Site 3, Site 8, and Site
9 (Fig. 2). The portion of metals distributed in the exchangeable
phase, an indication of the anthropogenic origin and of high potential bioavailability, were found to be low (<10%) for all metals and
for all sites. Fe or Mn oxides were important binding sites for Cd,
Zn, and Pb in the sediments from the three sites (30–60%). The
two oxides are important metal scavengers in sediments through
various mechanisms including coprecipitation, adsorption, surface
complex formation, ion exchange and penetration of the lattice
(Filgueiras et al., 2002). Cu was found to be mainly bound to the organic fraction in the samples collected from the fish market (S9).
This may be due to the high input of organic matters from the fish
market itself. Cu can easily form complexes with organic matter
due to the high stability constant of the organic–Cu complex. The
close association of Cu with the organic phase indicated the
anthropogenic origin of this metal (Fan et al., 2002b). A significant
portion (30–90%) of Ag, Cu, and Zn was also distributed in the residue fraction, indicating their principally non-anthropogenic origin. Although the Ag concentration in Site 9 was high, it
Table 2
Metal concentration in the Red Sea sediment samples (mean ± SD, mg kg1, except lg kg1 for Hg, dry weight basis, n = 3). Results are compared with crust metal background
concentrations and numerical sediment quality guidelines.
Site
Cd
Zn
Ag
Cu
Pb
As
Hg
S1
S2
S3
S4
S5
S6
S7
S8
S9
0.035 ± 0.007
0.024 ± 0.005
0.049 ± 0.009
0.027 ± 0.004
0.027 ± 0.002
0.029 ± 0.003
0.080 ± 0.021
0.110 ± 0.024
0.238 ± 0.024
10.0 ± 0.6
9.0 ± 0.6
45.5 ± 1.5
5.3 ± 0.4
4.9 ± 0.9
13.8 ± 0.4
46.6 ± 1.4
39.3 ± 1.3
179.0 ± 28.8
0.050 ± 0.003
0.067 ± 0.012
0.060 ± 0.002
0.069 ± 0.001
0.071 ± 0.001
0.068 ± 0.004
0.138 ± 0.002
0.387 ± 0.014
0.945 ± 0.054
0.69 ± 0.09
1.44 ± 0.09
18.48 ± 0.45
0.65 ± 0.20
0.45 ± 0.12
5.99 ± 0.23
8.69 ± 3.55
21.38 ± 0.66
82.99 ± 6.04
0.93 ± 0.15
2.74 ± 0.29
5.79 ± 0.18
0.49 ± 0.06
0.46 ± 0.06
1.52 ± 0.44
1.48 ± 0.21
2.72 ± 0.37
69.38 ± 7.55
1.5 ± 0.6
4.2 ± 0.3
6.3 ± 0.9
1.4 ± 0.8
2.1 ± 0.7
21.0 ± 7.7
2.2 ± 0.9
5.9 ± 1.1
2.5 ± 0.9
3.7 ± 0.9
40.4 ± 9.4
9.8 ± 0.7
3.2 ± 0.3
3.0 ± 0.3
4.6 ± 0.8
18.6 ± 4.2
50.4 ± 7.3
132.8 ± 20.0
0.098
0.99
3.0
5.0
71
120
290
460
0.05
1.6
1.9
2.2
25
32
91
150
20
36
83
130
1.5
9.8
21.4
33
20
180
640
1100
Obhor Sharm (outlet)
Marine station
Obhor Sharm (inner)
Abu Madafi reef
Non-reef reference site
Mangrove site
Treatment plant outfall
Entrance of fish market
Fish market
Crust concentrationa
SQG TEC
SQG MEC
SQG PEC
TEC, threshold effect concentration; MEC, median effect concentration; PEC, probable effect concentration; SQG, sediments quality guidelines.
a
(Taylor and Mclennan, 1995; Hare et al., 2010).
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K. Pan et al. / Marine Pollution Bulletin 62 (2011) 1140–1146
100
Cd
80
Exchangeable
Carbonate
60
Fe & Mn oxides
40
Organic matters
Residue
20
0
S3
S8
S9
100
% Extracted
Ag
Pb
80
60
40
20
0
S3
S8
S9
S3
S8
S9
S3
S8
S9
100
Zn
Cu
80
60
40
20
0
S3
S8
S9
Site
Site
Fig. 2. Relative distribution of metals in different geochemical species and in sediments.
appeared that 95% of the Ag was bound to the residue fraction.
Metals distributed in the residual phase are generally assumed to
have low mobility and bioavailability. Overall, the results of
sequential extraction further confirmed that the anthropogenic in-
put of metals was not significant in the studied areas except in the
fish market.
A total of nine species of sponges was collected from S1, S2, S4,
S5 and S6 as listed in Table 3. The pattern of the bioaccumulation of
Table 3
Metal concentration in sponge samples (mean ± SD, mg kg1, except lg kg1 for Hg, dry weight basis, n = 3).
Sponge species
Sampling location
Cd
Zn
Ag
Cu
Pb
As
Hg
Hyrtios erectus
S1
S2
S4
S5
S5
S1
S2
S4
S5
S1
S2
S1
S2
S2
S5
S6
S6
0.19 ± 0.05
0.23 ± 0.09
0.33 ± 0.03
0.56 ± 0.05
0.43 ± 0.05
0.17 ± 0.02
0.17 ± 0.07
0.22 ± 0.07
0.14 ± 0.05
0.75 ± 0.04
0.27 ± 0.03
0.16 ± 0.08
0.49 ± 0.11
0.23 ± 0.04
0.03 ± 0.02
0.69 ± 0.11
21.5 ± 5.10
24.3 ± 0.7
34.3 ± 3.2
7.7 ± 1.4
96.0 ± 15.9
27.1 ± 1.6
13.9 ± 1.3
18.0 ± 1.5
5.6 ± 0.6
47.5 ± 8.6
179.6 ± 6.0
200.7 ± 5.0
14.2 ± 1.2
19.5 ± 2.0
55.1 ± 6.3
86.7 ± 14.0
43.6 ± 5.5
154 ± 19.3
0.07 ± 0.02
0.04 ± 0.01
0.06 ± 0.01
0.22 ± 0.02
0.07 ± 0.01
0.05 ± 0.01
0.06 ± 0.01
0.06 ± 0.01
0.05 ± 0.01
0.06 ± 0.01
0.05 ± 0.01
0.06 ± 0.02
0.07 ± 0.02
0.08 ± 0.01
0.03 ± 0.01
0.09 ± 0.01
0.01 ± 0.01
23.9 ± 2.7
19.3 ± 4.4
18.4 ± 1.0
25.3 ± 1.3
6.7 ± 1.6
21.2 ± 1.1
16.3 ± 1.2
11.2 ± 0.9
22.5 ± 4.8
9.1 ± 0.7
15.5 ± 1.2
4.5 ± 1.1
6.7 ± 2.1
22.4 ± 2.1
8.3 ± 0.8
22.7 ± 2.3
8.0 ± 1.5
0.44 ± 0.02
0.29 ± 0.02
0.43 ± 0.06
0.35 ± 0.01
0.44 ± 016
0.56 ± 0.07
0.24 ± 0.02
0.36 ± 0.04
0.37 ± 0.23
0.73 ± 0.05
0.88 ± 0.10
0.28 ± 0.05
0.27 ± 0.04
2.07 ± 0.23
0.43 ± 0.01
1.79 ± 0.63
0.73 ± 0.22
23.5 ± 0.6
63.9 ± 7.0
15.0 ± 1.1
48.8 ± 4.9
2.3 ± 0.5
10.7 ± 4.8
8.1 ± 3.5
8.0 ± 3.8
7.5 ± 1.4
13.7 ± 2.2
22.1 ± 2.4
20.7 ± 2.1
42.2 ± 3.9
8.0 ± 2.2
11.2 ± 11.8
106.1 ± 6.5
15.0 ± 3.0
361 ± 17.9
360 ± 31.2
164 ± 20.7
384 ± 55.9
60.2 ± 7.0
74.6 ± 8.5
63.0 ± 10.1
55.5 ± 5.3
120 ± 4.4
108 ± 10.2
160 ± 7.0
231 ± 7.0
274 ± 20.2
29.8 ± 1.5
124 ± 3.5
266 ± 28.0
16.2 ± 0.6
Hyrtios sp.
Stylissa carteri
Chalinula sp.
Xestospongia testudinaria
Phyllospongia papyracea
Amphimedon sp.
Spongia arabica
Spheciospongia inconstans
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K. Pan et al. / Marine Pollution Bulletin 62 (2011) 1140–1146
metals varied among different species. Although the metal concentrations in the ambient environments were low, significant accumulation of Cd, Zn, Cu, and Hg was found in the sponges. In fact
the concentrations of metals in these sponges were several to hundreds of times higher than those in the sediments reported in previous studies (Patel et al., 1985; Philp, 1999; Cebrian et al., 2007).
For example, the concentration of Zn in Chalinula sp. was found to
be 180–200 lg g1, which was higher than the reported values
(100 lg g1) in some bivalve species such as clams and green
mussels (Blackmore, 1998). The strong ability to accumulate metals in sponges may be attributed to their high filtration rates of seawater (Hansen et al., 1995; Turon et al., 1997). For example,
Riisgärd et al. (1993) reported that the maximum pumping rate
for Halichondria panacea and Haliclona urceolus can be up to
86 L day1 g1. Both metals dissolved in water and adsorbed on
particles can be accumulated by sponges because sponges can retain up to 80% of suspended particles including the free-living bacteria particles and the size of the particles retained can be as small
as 0.028 lm (Reiswig, 1971; Milanese et al., 2003). In contrast, the
concentrations of Ag and Pb in sponges were found to be comparable to those in sediments, suggesting that the enrichment factors of
Ag and Pb were lower than those of the other metals. Indeed,
Cebrian et al. (2007) observed that sponges contained a low Pb
concentration even in an environment highly contaminated by
Pb, indicating that Pb was less bioaccumulated by sponges than
the other metals.
Distinct metal concentrations were found for different sponge
species collected from the same sites. For example, H. erectus generally had higher metal concentrations of Zn, As, Hg than S. carteri.
A typical concentration of around 300 ng g1 Hg was found for H.
erectus, which was higher than that for S. carteri (p < 0.01). Most
sponge species generally had relatively low Zn concentrations
(<50 lg g1 except for S. inconstans), but Chalinula sp. accumulated
up to 200 lg g1 of Zn which was several times higher than the levels accumulated by the other sponges sampled from the same sites
(p < 0.01). A high concentration of Cd (21.5 lg g1) was found in S.
inconstans. Interestingly, Zn was also highly accumulated in this
sponge (154 lg g1). The enrichment factor of Cd (calculated as
the ratio of metal concentration to that of sediments) for S. inconstans was 700, which was about 70 times that of other sponge species. The background level of dissolved Cd in the pristine sites was
reported to be 20–40 ng L1 (Hall et al., 1996; Jensen and BroRasmussen, 1992). The bioconcentration factor of Cd for S. inconstans reached 106 L kg1, indicating the sponge’s special preference
for Cd. Another species of sponge S. arabica accumulated relatively
high concentrations of As and Hg (106 lg g1 and 265 ng g1,
respectively). However, the reason why the sponges accumulated
such high concentrations of specific metals remains unclear.
The interspecies variations in metal concentrations observed in
the sponges in this study are not uncommon. Patel et al. (1985)
measured the concentrations of 17 trace metals in two species of
sponges, Spirastrella scupidifera and Prostylyssa foetida, and found
that the former accumulated significantly higher Ni (400–
2250 lg g1) than the latter (7–15 lg g1). It was possible that
the interspecies difference may reflect the dissimilarity of pumping
physiology in various species, such as the volume of choanocyte
chambers (Cebrian et al., 2007). Other studies attributed the variation to the influences of environmental conditions such as pH,
salinity and geographical features (Philp, 1999; Bargagli et al.,
1996). Apart from the above factors, it is possible that the array
of microorganisms attached to the sponges may also contribute
importantly to the metal accumulation, an area remains rather
unexplored in previous studies. In this study, there was actually
contrasting difference in the microbial community structure for
H. erectus and S. carteri. Proteobacteria, Firmicutes and Chloroflexi
altogether constitute up to 52–73% of the community attached to
H. erectus, while Proteobacteria was the major phylum in S. carteri
and made up 60% of the community (Lee et al., 2010). Such specific
difference in the microbial community structure may also be
responsible for the different accumulation patterns of Zn, As, and
Hg between the two species. Sponges are ecologically diverse hotspots of unexplored microbial communities. Bacteria, unicellular
algae, cyanobacteria, dinoflagellates, zoochlorellae and domain
Archaea have been found to attach themselves to the sponges in
the extra- and intra-cellular spaces (Lee et al., 2009). The presence
of sponge-associated bacteria can have significant impact on the
metal concentrations in marine sponges, because they can account
for up to 40–50% of a sponge’s biomass (Hentschel et al., 2002;
Selvin et al., 2009). On the one hand, bacteria are efficient metal
bioaccumulators that either adsorb or absorb metals because of
their high surface-volume ratio (Dixon et al., 2006; Chen et al.,
2008). It has been shown that the accumulation of metals is also
species-specific for bacteria (Vogel and Fisher, 2010). On the other
hand, bacteria can facilitate metal uptake in the organisms they are
attached to. For example, Sayler et al. (1975) found that mercury
concentrations were 200 times greater in tissue fractions of the
oyster Crassostrea virginica dosed with the mercury-metabolizing
bacteria Pseudomonas compared with oysters not dosed with the
bacteria in the control. Therefore, it is reasonable to think that
the specific bacterial communities attached to the sponges can
affect the interspecies metal accumulation patterns of their hosts.
Currently there are about 240 known species of sponge in the
Red Sea, with many species uninvestigated (Ilan et al., 2004;
Radwan et al., 2010). Because of their abundant resources and
strong capacities for metal accumulation, sponges appear to be
promising biomonitors for the Red Sea coastal environment.
Sponges show a wide distribution in coastal environments, and
are available all year round, are abundant in sublittoral areas,
and are easy to transplant (Cebrian et al., 2007). They are sedentary
benthic invertebrates which means that the accumulated contaminants detected are representative of the ones found in the area.
The concentrated metals in sponge tissues allow more simple measurements of metals to be carried out than the critical techniques
required for the typical water analysis. Our results show that
sponges had species-specific affinities for metals, which suggests
that biomonitoring data for a certain contaminant should be compared intraspecifically. Attention should be given to the accumulation strategy when employing sponges as biomonitors. A strong
net accumulator is better than a weak one for monitoring programmes. For example, it may be more appropriate to use S. inconstans and Chalinula sp. to monitor Cd and Zn contamination,
respectively, because they can accumulate the respective metals
readily. Employing multiple species of sponges may generate complementary results.
A biomonitor should be able to accumulate a considerable concentration of contaminant in relation to the average contaminant
concentration in its ambient environment. Cebrian et al. (2007)
found significant positive relationships between the metals accumulated in the sediments and in the sponges found in Mediterranean coasts. Such relationships were not found in our study
possibly because only a limited number of sponge samples were
available from each sampling site, and the metal exposures among
different sites in which sponges were collected were similarly low.
However, it is noted that the accumulation of metals in sponges occurs both directly from dissolved metals via water filtration, as
well as from particulate metals when food particles are retained
by the sponges (Roberts et al., 2008). Metal concentrations in the
sediment provide valuable information about metal contamination
in the ambient environment but do not supply comparative information about the metals’ bioavailability to sponges. Beside metal
concentrations in ambient environments, geographical features
such as water currents may also affect the accumulation of metals
1145
K. Pan et al. / Marine Pollution Bulletin 62 (2011) 1140–1146
60
40
y=0.47x+2.0
2
r =0.99
Metal concentration in S.carteri (µg g-1)
45
y=1.4x-13.6
2
r =0.80
30
p <0.05
p <0.01
30
20
15
10
Zn
Cu
0
0
0
30
60
90
120
0
0.8
0.20
0.6
0.15
0.4
0.10
0.2
10
20
30
40
0.05
Pb
Hg
0.0
0.00
0.0
0.2
0.4
0.6
0.8
0.1
0.2
0.3
0.4
0.5
Metal concentration in H.erectus (µg g-1)
Fig. 3. Relationship between the metal concentrations in the sponges H. erectus and S. carteri collected concurrently from four different sites (n = 3).
in sponges. Bargagli et al. (1996) suggested that intense upwelling
currents were responsible for a high accumulation of cadmium by
Antarctic sponges grown in the pristine Antarctic areas. Uncertainties remain about the contribution of dissolved metals and the effects of other geographical factors on the overall metal
accumulation in sponges in our study. Interestingly, significant
relationships were found between the Zn and Cu accumulated in
H. erectus and S. carteri collected from different sites (Fig. 3), implying that there was a positive relationship between the metal concentrations in the sponges and the bioavailability of Zn and Cu at
these sites. The results also suggest the feasibility of employing
multi-species of sponges in biomonitoring to provide more reliable
and consistent results.
The metal concentrations in sediments collected from the nine
sampling sites were low. Both the total concentration and the geochemical speciation of metals showed that metal contamination
was not significant in most of the studied areas. Only the fish market and its adjacent areas were moderately metal contaminated.
Sponges accumulated much higher metal contents than the
sediments although the metal exposure was low in the ambient
environment. Significant interspecies variations in the metal
accumulation patterns were found in the sponges. Sponges in the
Red Sea appear to be promising biomonitors of metal contamination because of their excellent ability to accumulate specific
metals.
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