Effective metal concentrations in porewater and seawater labile

Marine Pollution Bulletin 44 (2002) 956–976
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Edited by Bruce J. Richardson
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Effective metal concentrations in porewater and seawater labile
metal concentrations associated with copper mine tailings
disposal into the coastal waters of the Atacama region of
northern Chile
M.R. Lee
a
a,*
, J.A. Correa a, H. Zhang
Departamento de Ecologıa and Center for Advanced Studies in Ecology and Biodiversity, Facultad de Ciencias Biol
ogicas,
Pontificia Universidad Catolica de Chile, Casilla 114-D, Santiago, Chile
b
Department of Environmental Science, IENS, Lancaster University, Lancaster LA1 4YQ, UK
Tailings dumping on the coast of the Atacama region
takes place in the area around the small city of
Cha~
naral. The coast of this area is open to the Pacific
Ocean swell and consists of rocky shores interspersed
with sandy beaches of coarse sediments. Freshwater in
this area is a scarce resource and significant freshwater
input to the coastal environment occurs in rare flash
floods only. As a result of the desert nature of the area
population density is low, there is no agricultural activity and the only industrial activity is mining. Therefore, the only significant source of pollution in this
costal environment is that derived from the mining activities, providing a unique opportunity to study the
effects of tailings disposal and elevated metal concentration in the absence of the confounding effects of other
groups of pollutants.
Mine tailings from the copper mines at Potrerillos
(now closed) and El Salvador have been dumped into
coastal waters since 1939. Initially the tailings were deposited on to the beach at Playa Cha~
naral (Fig. 1) and
as a result the shore line is now a kilometre further out
to sea than it had been originally (Castilla, 1983). In the
*
b
Corresponding author. Fax: +56-35-431670.
E-mail address: mlee@genes.bio.puc.cl (M.R. Lee).
1970s the tailings were diverted to a new dumping point
ten kilometres to the north of Cha~
naral, adjacent to
Playa Palito (Fig. 1). Again a large tailings beach
formed in the littoral zone between Caleta La Lancha
and Caleta Agua Hedionda, with significant amounts of
tailings mixed in with the natural sediments on the
beaches around Caleta Palito. In 1990, after local court
action, a tailings settlement dam was constructed inland
at Pampa Austral (60 km east of Cha~
naral) and since
then only ‘clear water’ tailings have been delivered to the
coast with a legal maximum concentration of 2000 lg
Cu l1 (Castilla, 1996).
The data included in this report were collected as part
of a larger study on the effects of copper on the coastal
environment and specifically as part of a sediment
quality triad. The objective of the latter was to understand the effects of the tailings disposal on the littoral
meiofaunal communities of the area. The sampling sites
were divided a priori into three groups: the northern
sites, Puerto Pan de Azucar (Pue), Frente Isla Pan de
Azucar (Fre), Playa Blanca (Bla); the central ‘impacted’
sites, Caleta La Lancha (Lan), Caleta Agua Hedionda
(Hed), Palito 1000 m Norte (Mil), Playa Palito (Pal),
Palito 2000 m Sur (Dos), Playa Cha~
naral (Cha); and the
reference sites, Las Piscinas (Pis), Torres del Inca (Tor),
Playa Zenteno (Zen). See Fig. 1 for the locations of the
Baseline / Marine Pollution Bulletin 44 (2002) 956–976
957
Fig. 1. A map of the Cha~
naral area showing the locations of the sampling sites and the tailings dumping point.
sampling sites, all of which will here after, in figures and
tables, be referred to by their three letter codes.
The effective concentrations of metals in the sediment
porewater, August 1999, and the concentrations of labile metals in the overlying seawater, March 2000, were
determined using the diffusive gradients in thin-film
(DGT) technique (Davidson and Zhang, 1994; Zhang
and Davidson, 1995; Zhang et al., 2001).
The technique of DGT has been pioneered and developed at Lancaster to measure metals and to study
their chemical speciation in natural waters (Davidson
and Zhang, 1994; Zhang and Davidson, 1995, 2000) and
their effective concentrations in soils and sediments
(Zhang et al., 2001). It accumulates solutes quantitatively on a binding agent after their passage through a
well-defined diffusion layer. A polyacrylamide hydrogel,
of known thickness (Dg), is commonly used as the dif-
fusive layer, while for trace metals Chelex-100, incorporated into a second gel layer, serves as the binding
agent. The two gels are enclosed in a small plastic device
that is immersed in solution. The diffusion coefficient
(D) of metal ions within the diffusive gel can be independently measured. Fick’s first law of diffusion can
then be used to determine the concentration (Cb ) of
metal in the bulk solution from the mass (M) of metal
accumulated in the resin, as shown in Eq. (1).
Cb ¼ MDg=ðDtAÞ
ð1Þ
The DGT device provides a direct measurement of the
effective concentration CE in sediment (Zhang et al.,
2001). When the DGT device is deployed in sediment it
locally lowers the concentration of metal ions and induces the re-supply from sediment to porewater in the
same manner as biological organisms. It measures the
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Baseline / Marine Pollution Bulletin 44 (2002) 956–976
concentration from the porewater plus the effective additional concentration supplied from the solid phase.
This effective concentration, CE , is what would be experienced by the organisms living in the sediments under
the worst possible case, i.e. when the membrane transport is rapid.
The sampler consists of a plastic base and cap, which
are used to hold the gel ‘sandwich’ in place. The base
layer of the ‘sandwich’ is a gel impregnated with Chelex
beads, which bind the labile metals effectively. The second layer is the diffusion gel, with a large enough pore
size to allow metal ions and complexes to diffuse
through. The third layer is a 0.45 lm cellulose-nitrate
filter that protects the gel surface from mechanical
damage. The gels and plastic holders were supplied by
DGT Research Ltd. UK (www.dgtresearch.com).
The DGT samplers, including three blanks, were assembled less than a week before they were due to be
used. The plastic bases and caps were acid washed (10%
HNO3 for 48 h) and then rinsed in ultrapurTM water
prior to assembly. The samplers were assembled in a
laminar flow chamber to prevent contamination and
stored at 4 °C, with a small amount of ultrapurTM water
to prevent desiccation, in plastic bags prior to use.
Samplers were transported to and from the study site in
coolers. Due to the dynamic nature of the littoral environment on this part of the Chilean coast it was not
possible to leave the samplers in situ. Therefore, for the
porewater samples three replicate cores were collected
from each site. The cores were collected using PVC
tubing 45 mm in diameter, 150 mm in length, which had
been acid washed prior to use (10% HNO3 for 48 h). The
core tubes were inserted into the sediment to a depth of
approximately 50 mm, removed with the sample and a
cap placed on the top of the tube. The sample was then
inverted and a DGT unit placed face down in the sediment. The cores were then placed in ZiplocTM bags and
then into a cooler. Seawater samples (three replicate
samples for each site) were collected from the surf zone
on each beach. Water was collected from the surf using a
plastic bucket. One litre of seawater was placed into a
ZiplocTM bag along with a DGT unit. The bag was then
placed in a second ZiplocTM bag for added security/
durability and the bags placed in a cooler. The exposure
time of the DGT sampler to the water was at least 24 h
for each sample. However, exposure times varied from
site to site due to logistical constraints. At the end of the
exposure, the DGT units were removed from the samples, rinsed in ultrapurTM water, placed in clean plastic
bags and stored in a cooler for the return journey to the
laboratory.
In the laboratory the Chelex gels were removed and
placed into 1.5 ml micro-centrifuge tubes with 0.8 ml of
10% HNO3 (Merck suprapur) and then sent by courier
to the University of Lancaster for analysis. Samples
were analysed by inductively coupled plasma-mass
Table 1
The results of ANOVA tests on the effective metals concentrations
data for porewater and labile metal concentrations for seawater, collected from 12 sites in the Cha~
naral area
Sample
Metal
F
p
Porewater
Cu
Al
As
Cr
Mn
Ni
Zn
4.806
1.880
8.958
9.526
0.963
52.385
2.542
0.001
0.095
0.0001
0.0001
0.503
0.0001
0.027
Seawater
Cu
Al
As
Cr
Mn
Ni
Zn
45.106
6.145
2.463
9.737
4.159
1.577
16.606
0.0001
0.0001
0.031
0.0001
0.002
0.169
0.0001
spectrometry (ICP-MS, Varian Ultramass) using a direct injection nebulizer (CETAC). During the analyses a
quality control solution was used for every 10 samples to
monitor the performance of the machine. The blanks,
which included all steps in the sampling process––
prepartion of the DGT device, handling and analysis,
had concentrations of less than 2 ng l1 for all the metals
studied.
The effective concentrations in the porewater and the
labile concentrations in the seawater were quantified for
the following metals: Cu, Al, As, Cr, Mn, Ni, and Zn.
The subscripts pw and sw refer to porewater and seawater respectively. The data for all the metals, both
porewater and seawater, were normally distributed.
However, Cupw , Cusw , Mnpw , Nisw , and Znsw all needed
to be log transformed to achieve normal distribution.
Analysis of variance (ANOVA) tests (Table 1) for each
of the metals, both in porewater and seawater, indicated
that all metals showed significant (p < 0:05) between-site
variation except Alpw , Mnpw , and Nisw .
The effective concentrations of metals in the porewater were consistently higher than the labile concentrations measured for the seawater. Seawater
concentrations of copper, aluminium, chromium, manganese, and zinc can be visually associated with the expected distribution of tailings impact. However, for
concentrations in the porewater, this only applies to
copper. Only in the case of copper did the seawater and
porewater concentrations vary in the same way between
sites (Fig. 2). The maximum and minimum concentrations of labile metals in both the porewater and the
seawater for each of the metals are presented in Table 2.
The concentrations of copper in both the porewater
(effective) and seawater (labile) were higher in the central
sites than they were at the northern and reference sites
(Fig. 2a). Copper concentrations were also higher at the
northern sites than at the reference sites. Effective
Baseline / Marine Pollution Bulletin 44 (2002) 956–976
959
Fig. 2. The effective porewater metal concentrations and seawater labile metal concentrations (lg l1 ) recorded at each of the sites (bars represent
standard errors).
Table 2
The maximum and minimum effective concentrations in porewater and labile concentrations in seawater for each of the metals (lg l1 ) collected from
12 sites in the Cha~
naral area
Porewater
Copper
Aluminium
Arsenic
Chromium
Manganese
Nickel
Zinc
Seawater
Max
Min
Max
Min
1449.59 (Hed)
820.92 (Bla)
6.14 (Hed)
15.08 (Fre)
12.38 (Pal)
43.71 (Pis)
1263.18 (Pis)
6.43 (Pis)
179.78 (Zen)
1.82 (Zen)
3.75 (Zen)
3.17 (Zen)
18.01 (Cha)
493.52 (Zen)
41.42 (Hed)
189.48 (Mil)
1.22 (Hed)
2.94 (Cha)
4.24 (Hed)
3.11 (Fre)
631.13 (Cha)
0.16 (Pis)
34.87 (Pue)
0.13 (Lan)
1.05 (Pue)
1.22 (Zen)
1.33 (Tor)
225.18 (Tor)
porewater concentrations of aluminium did not vary
significantly between sites (Fig. 2b). In seawater, however, the concentrations of labile aluminium tended be
higher at the sites closer to Caleta Palito. Seawater
concentrations of labile arsenic fluctuated from site to
site with no apparent pattern emerging (Fig. 2c). On
the other hand, effective arsenic concentrations in the
porewater fluctuated little, but with a peak at Caleta
Agua Hedionda. The speciation of arsenic is currently
under investigation using the DGT technique. Preliminary results indicated that the DGT device used in this
work (using Chelex-100 as binding agent) only measure
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Baseline / Marine Pollution Bulletin 44 (2002) 956–976
Fig. 3. A plot of effective porewater metal concentration versus seawater labile copper concentration with regression analysis. The data was logðx þ 1Þ
transformed prior to plotting.
As3þ . Effective porewater chromium concentrations did
not vary much between sites (Fig. 2d) although moving
northwards, concentrations tended to increase. Concentrations of chromium in the seawater were highest at
the central impacted sites, with a steady decline in
concentration moving northwards from Playa Cha~
naral.
The chromium measured by DGT (with Chelex-100 resin) is Cr3þ only (Ernstberger et al., 2000). The effective
porewater concentrations of manganese did not vary
significantly between sites (Fig. 2e) whereas labile concentrations in the seawater were lowest at the reference
sites. The nickel concentrations in both the porewater
(effective) and the seawater (labile) did not appear to
vary much between sites (Fig. 2f), except for a peak
value recorded in the porewater at Las Piscinas. Effective porewater zinc concentrations did not show much
variation between sites (Fig. 2g). However, the reference
sites clearly had lower labile concentrations in the
seawater, with the peak concentration occurring at
Playa Cha~
naral with a steady decline in concentrations
moving northwards.
Of the Pearsons correlations between the porewater
(effective) and seawater (labile) concentrations of a given
metal, only copper had a significant correlation (0.596,
p < 0:001). Regression analysis of porewater (effective)
versus seawater (labile) copper concentrations (Fig. 3)
indicated that there was a significant positive relationship between the two (F ¼ 55:283, p < 0:0001).
Regression analysis yielded the following equation:
y ¼ 0:577x 0:033, R2 ¼ 0:847 (where x is the effective
porewater concentration and y the labile seawater concentration) indicating that the concentration of labile
copper in the seawater increased with the effective concentration of copper in the porewater.
Spearman’s rank correlation tests were carried out
with the metal concentrations in both the porewater
(effective) and the seawater (labile), and a ranking of the
sites based on the amount of tailings present. Rankings
by the amount of tailings present (Table 3) were based
upon the visual observation of the sites outlined above,
and were therefore subjective. The sites were ordered so
that the sites with least amount of tailings receiving the
lowest rank. Caleta La Lancha (Lan), Caleta Agua
Hedionda (Hed) and Playa Cha~
naral (Cha) are all sites
composed of 100% tailings. The results of the Spearman’s rank test are presented in Table 4 and show that
there was a high correlation between the tailings and
Table 3
The study sites ranked by the amount of tailings present at each site,
based on visual observation
Site
Ranking
Pue
Fre
Bla
Lan
Hed
Mil
Pal
Dos
Cha
Pis
Tor
Zen
6
5
4
11
11
8.5
7
8.5
11
2
2
2
Sites with the least amount of tailings were given the lowest ranks.
Table 4
Spearman’s rank correlation of tailings and effective metal concentrations in porewater and labile metal concentrations in seawater from
12 sites in the Cha~
naral area
Metal
Seawater
Porewater
Cu
Al
As
Cr
Mn
Ni
Zn
0.913
0.405
0.146
0.430
0.799
0.732
0.831
0.913
0.135
0.444
0.245
0.025
0.345
0.242
Baseline / Marine Pollution Bulletin 44 (2002) 956–976
copper in both the porewater (0.913) and the seawater
(0.913). In the porewater, only the effective copper
concentration showed a high correlation with the tailings, with all other metals displaying low correlations
(<0.500). In the seawater samples in addition to copper,
manganese (0.799), nickel (0.732) and zinc (0.831)
showed high correlations with the tailings. The other
metals in the seawater had low correlations (<0.500)
with the tailings.
Copper is the principal component of past tailings
discharges from the perspective of sediment toxicology. Only Cupw and Cusw are highly correlated with
the distribution of the tailings and with each other. The
porewater copper concentrations recorded here are considerably higher than those recorded by Alongi et al.
(1991) for the Fly River Delta, Papua New Guinea.
Alongi et al. (1991) is the only other study, to the authors knowledge, that reports the porewater copper
concentrations in tailings discharges. Manganese, nickel
and zinc were all highly correlated with the observed
distribution of the tailings but only in the seawater.
The metals recorded in the area under study can be
divided into two groups: those present at natural background concentrations and those whose concentrations
have been increased at some of the sites by anthropogenic sources. The metals present at natural background
levels are characterised by concentrations that either
fluctuate little between sites or appear to fluctuate in a
pattern that can not be tied in with current knowledge of
the anthropogenic sources, (i.e. at random). Within the
porewater, only the effective concentrations of copper
appear to be above background levels; therefore, all the
other metals sampled are present, within the porewater,
at natural background concentrations. For the seawater
samples, the metals can be divided into four groups. The
first group consists of arsenic and nickel which are
present at background concentrations only. The second
group of metals includes aluminium and manganese,
which displayed elevated concentrations around the
tailings discharge point at Caleta Palito and are probably associated with the ‘clear water’ tailings discharge.
The third group includes chromium and zinc, where the
highest concentrations occur at Playa Cha~
naral with a
steady decline in concentration moving northward. The
only obvious possible source of metals near Playa
Cha~
naral is the port activity at Barquito located one
kilometre to the south of Cha~
naral. The final, singlemetal group includes only copper, where the source of
the metal in the seawater is clearly the tailings that have
been deposited in the intertidal zone.
In summary, of the metals reported in the present
study, only copper appears to be ecotoxicologically
important in the porewater. The tailings themselves, in
addition to their physical impact, are also the primary
sources of copper in the seawater adjacent to the sites.
A comparison of the data presented here with data from
961
other sites is difficult as in the majority of cases the
concentrations of metals within sediments are measured
in the form of lg g1 DW of sediment. This measure of
metals in not ecologically relevant from the point of
view of ecotoxicology. It is well known that the bulk
sediment concentrations of a metal are not predictive of
the toxic effect of those sediments (Luoma, 1989; Di
Toro et al., 1990; Chapman et al., 1998). When only
bulk sediment concentrations are known, a host of other
variables need to be measured and applied to complex
models in order to estimate the bioavailable fraction of
the metals. Therefore, it would seem more appropriate
in the future to measure the biologically more appropriate fractions of the metals present using a methodology such as DGT.
Acknowledgements
The authors thank the International Copper Association for providing research funds for this study.
CASEB, program 7, is also acknowledged.
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