Download Report

MARINE
ENVIRONMENTAL
RESEARCH
Marine Environmental Research 59 (2005) 1–18
www.elsevier.com/locate/marenvrev
Eﬀects of copper mine tailings disposal on littoral
meiofaunal assemblages in the Atacama region
of northern Chile
Matthew R. Lee *, Juan A. Correa
Departamento de Ecologıa and Center for Advanced Studies in Ecology and Biodiversity, Facultad de
Ciencias Biologicas, Pontiﬁcia Universidad Catolica de Chile, Alameda 340, Santiago, Chile
Received 10 October 2003; accepted 20 January 2004
Abstract
The eﬀects of the disposal of copper mine tailings on the littoral meiofaunal assemblages of
the Cha~
naral area of northern Chile were studied. Of the metals data collected, only in the case
of copper was there a clear association with the tailings distribution in both the seawater and
porewater samples, and it is assumed that the tailings on the beaches was the source of copper
in the adjacent seawater. When compared to the reference sites, the meiofaunal assemblages at
the impacted sites had signiﬁcantly lower densities and taxa diversities; at the northern sites
only the densities were lower. Otoplanid turbellarians were identiﬁed as characteristic of those
beaches impacted by tailings. The combination of porewater copper and the amount of tailings present were identiﬁed as mostly responsible for the observed structure of the meiofaunal
assemblages. It was also established that the variation in natural sediment grain size from
beach to beach was not a signiﬁcant factor in the observed diﬀerences in the meiofaunal assemblages. The two groups of meiofauna that proved to be most sensitive to the eﬀects of
tailings dumping were the foraminiferans and the harpacticoid copepods.
Ó 2004 Elsevier Ltd. All rights reserved.
Keywords: Meiofauna; High-energy sandy beaches; Tailings; Metals; Copper; Chile
*
Corresponding author. Present address: 11 Briar Drive Heswall, Wirral, Merseyside CH60 5RW, UK.
E-mail address: leemr@btopenworld.com (M.R. Lee).
0141-1136/$ - see front matter Ó 2004 Elsevier Ltd. All rights reserved.
doi:10.1016/j.marenvres.2004.01.002
2
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
1. Introduction
The eﬀects of contaminating discharges on marine ecosystems are studied, in most
industrialised countries, around estuarine or harbour environments and the chosen
organisms are frequently the abundant sediment dwelling macrofauna (e.g. Boening,
1999; Gren, Destouni, & Scharin, 2000; Stark, 1998). The discharge of contaminants
aﬀecting high-energy sandy beaches is less common and consequently, studies of the
eﬀects on the resident biota of this environment are scarce (Barros, 2001; Castilla,
1983; Watling & Watling, 1983). The nature of the sediments in estuarine and highenergy sandy beaches are signiﬁcantly diﬀerent. In the estuarine environment the
sediments are usually ﬁne with high concentrations of organic material and low
oxygen regimes, where as, in high-energy sandy beaches they are coarser, with low
concentrations of organic material and high oxygen regimes (McLachlan, 1983;
McLachlan & Turner, 1994). The importance of these features is considerable, as
they have a signiﬁcant eﬀect in determining both, the resident biota and the availability of the contaminants to the biota (Chapman, Wang, Janssen, Persoone, &
Allen, 1998). It is therefore highly questionable to take the information gained by
studies of the eﬀects of contaminants on the estuarine sedimentary environment and
apply it to the high-energy sandy beach environment. To address this problem
speciﬁc studies examining the eﬀects of contaminants on high-energy sandy beaches
are required.
When compared to the low-energy sedimentary environment, the macrofaunal
diversity of Chilean high-energy sandy beaches is low, typically between one and ten
species (Jaramillo, 1994; Jaramillo, McLachlan, & Coetzee, 1993) depending on the
physical nature of the beach. These macrofaunal species also have a pelagic dispersal
phase in their life-cycles which means that larvae, usually the most sensitive stage,
are not exposed to the contaminants in the sediment. This, combined with recruitment derived from populations outside the aﬀected area, can serve to mask the
impact that the contaminants have on the littoral environment. In this study we
therefore chose to use the meiofauna which are highly diverse and do not, as a rule,
have a pelagic dispersal phase. The diversity of meiofauna present provides a wide
range of physiological responses and, therefore, a better understanding of the true
impact of the contaminant. The many other advantages of meiofauna in these types
of studies are reviewed by Coull and Chandler (1992).
The situation in the Cha~
naral area (26°20:50 S, 070°37:40 W) of northern Chile
provides an ideal situation for a natural experiment on the impact of tailings
dumping in isolation. Due to the desert nature of this area other forms of contamination, such as industrial and agricultural eﬄuents, are absent. The dumping of
copper mine tailings into the coastal environment around Cha~
naral has taken place
since 1938 (Castilla & Correa, 1997). Initially the untreated tailings were dumped
into Cha~
naral Bay where a large beach composed entirely of tailings formed, the
dotted line in Fig. 1 indicates the approximate position of the original coastline. In
the mid 1970s the tailings were diverted to a new dumping point, approximately ten
kilometres to the north of Cha~
naral, again a large tailings beach formed north of the
dumping point. In 1990 a tailings settlement dam was constructed inland where the
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
3
Fig. 1. A map of Chile indicating Cha~
naral (a) and showing the locations of the sampling sites (b) the
northern and central sites, (c) the reference sites. The dashed line in (b) indicates the approximate position
of the original coastline.
solid component of the tailings is allowed to settle out before the tailings water, ‘clear
water’ tailings, continues on to the coast.
The objective of this study was therefore to examine the eﬀects of the copper mine
tailings disposal on the littoral meiofaunal assemblages of the area and to identify
those meiofaunal groups which could act as indicators of copper pollution.
2. Material and methods
Samples were collected from sites in the Cha~
naral area of northern Chile on ten
occasions between January 1997 and March 2000. Five replicate samples, each of 50
cm3 in volume (depth 75 mm), were collected from each site using a 60 cm3 plastic
4
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
syringe modiﬁed to form a quantitative piston corer. Samples were collected from
the surf zone on the lower portion of the beach, the zone of retention, at or around
low tide. Collection was at random and parallel to the shore line. Each sample was
placed in a 100 cm3 plastic bottle with approximately 50 cm3 of 10% formalin solution, shaken, sealed and then returned to the laboratory for processing. Samples
were collected from twelve sites covering the range of sediment and morphodynamic
types occurring in the area (See Table 1 for site descriptions, and Fig. 1 for locations). These sites were subdivided a priori into three groups. The ﬁrst group were
the reference sites Playa Zenteno, Torres del Inca and Las Piscinas (Fig. 1(c)), these
sites were assumed to be suﬃciently far south (100–150 km) as to be unaﬀected by
the tailings dumping. The second group were the northern sites Puerto Pan de
Azucar, Frente Isla Pan de Azucar and Playa Blanca, located within the Parque
Nacional Pan de Azucar and showed no evidence of tailings deposition. Finally, the
third group were the central sites Caleta La Lancha, Caleta Agua Hedionda, Palito
1000 m Norte, Playa Palito, Palito 2000 m Sur and Playa Cha~
naral, these sites are
centred around the dumping point at Caleta Palito and had varying degrees of
tailings deposition.
The eﬀective porewater (Zhang, Zhao, Sun, Davidson, & McGrath, 2001), and
the seawater labile (Davidson & Zhang, 1994), metal concentrations were measured
using the diﬀusion gel technique (DGT). The sampler consists of a plastic base and
cap which are used to hold the gel ‘sandwich’ in place. The base layer of the gel
‘sandwich’ is a gel impregnated with Chelex beads, which bind the labile metals
reaching them. The second layer is the diﬀusion gel, the pore size of which is designed
to allow only labile metal ions to pass. The third layer is a 0.45 lm cellulose-nitrate
ﬁlter which protects the diﬀusion gel from abrasion. The samplers were assembled in
a laminar ﬂow hood less than a week before they were due to be used, the plastic
parts were acid washed (10% HNO3 for 48 h) and then rinsed in ultrapurTM water
prior to assembly.
Table 1
The detailed location of the beaches used in this study
Beach
Puerto Pan de Azucar
Frente Isla Pan de Azucar
Playa Blanca
Caleta La Lancha
Caleta Agua Hedionda
Palito 100 m Norte
Playa Palito
Palito 2000 m Sur
Playa Cha~
naral
Las Piscinas
Torres del Inca
Playa Zenteno
Code
Md (u)
Pue
Fre
Bla
Lan
Hed
Mil
Pal
Dos
Cha
Pis
Tor
Zen
)0.03
1.54
2.43
1.96
1.94
)0.78
)0.78
)0.14
2.30
1.72
)0.10
1.67
Lat.
Long.
0
26°08:3 S
26°08:40 S
26°11:10 S
26°13:40 S
26°15:30 S
26°16:10 S
26°16:30 S
26°17:00 S
26°20:50 S
26°33:00 S
26°36:20 S
26°51:10 S
The column Md refers to the graphic mean sediment grain size in units of phi (u).
070°39:30 W
070°40:00 W
070°39:20 W
070°39:20 W
070°38:50 W
070°39:20 W
070°39:30 W
070°29:60 W
070°37:40 W
070°41:10 W
070°44:50 W
070°48:30 W
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
5
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 and 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 75 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 new ZiplocTM bags in a
cooler. The exposure time of the DGT sampler to the porewater 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, and placed in clean plastic bags for the return
journey to the laboratory in a cooler.
Water 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 new ZiplocTM bag along with a DGT unit. The
bag was then placed in a second new ZiplocTM bag for added security/durability and
the bags placed in a cooler. Again the minimum exposure time of the DGT unit to
the seawater was 24 h for each sample.
In the laboratory the Chelex gels were removed from the DGT units 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 spectrometry (ICP-MS, Varian Ultramass) using a direct injection nebulizer (CETAC). The detailed methodology
has been published elsewhere (Zhang, Davidson, Knight, & McGrath, 1998; Zhang
et al., 2001).
Analysis of the sediment structure was conducted using sediment from the meiofauna samples. Three of the ﬁve samples, from each site were selected at random for
sediment analysis. The samples were ﬁrst washed with fresh water on a 63 lm screen
to remove salt crystals, there was no silt component to the sediments at any of the
sites. The sediment samples were then dried for 72 h at 80 °C in a drying oven. The
dried sediment samples were then sieved through a stack of standard brass analytical
sieves, shaken for 15 min. Each fraction was weighed to the nearest 0.1 g. The following measures were calculated using phi (u) values, the graphic mean (Md, Table
1), the inclusive graphic quartile deviation (QDI) and the inclusive graphic skewness
(SkI). None of the measures was able to identify either the presence or proportion of
tailings present at a given site. Therefore, a qualitative assessment of the amount of
tailings present at each site was made based on visual observation (Table 6), the
tailings have distinct yellow colour which allows them to be distinguished from the
natural ﬁne sediments which are grey. Sites with the least amount of tailings were
given the lowest rank.
In the laboratory the meiofauna were extracted from the sand using a simple
decantation technique (Pfannkuche & Thiel, 1988). Samples were shaken for approximately 30 s and the supernatant poured through a 44 lm mesh. This process
was repeated ﬁve times for each sample (95% extraction eﬃciency, Lee, 2001). The
6
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
meiofauna were then washed into a girded petri dish and counted using a binocular
microscope (Wild M5, 200). The meiofauna were analysed to a higher taxonomic
resolution than species, typically orders, referred to herein as ‘taxa diversity’. The
underlying hypothesis for analysing samples at higher taxonomic resolution is that
species level analyses are more aﬀected by natural variation in environmental variables, such as sediment grain size, than are analyses at the level of family, order or
phylum (Warwick, 1988a). Furthermore, several studies have concluded that there is
no substantial loss of information when considering the overall eﬀects of a pollutant
on the environment when the biota is analysed at higher levels of taxonomic resolution, particularly where the pollution gradient is marked (Somerﬁeld & Clarke,
1995; Warwick, 1988b).
The meiofauna data collected were analysed using multivariate statistical analyses
(PRIMER, Plymsolve). The following multivariate analyses were conducted: analysis of similarities (ANOSIM), non-metric multidimensional scaling (MDS), similarities percentage analysis (SIMPER) and biotic/environmental variable analysis
(BIOENV).
3. Results
Of the metals analysed in this study only the copper concentrations appeared to
be associated with the distribution of the tailings so only they will be considered here.
The data and discussion for all the metals sampled in this study is presented elsewhere (Lee, Correa, & Zhang, 2002). The highest concentrations of copper were
associated with those beaches that had solid tailings waste present. The pattern of
eﬀective porewater copper distribution in the sediment was mirrored by that of the
labile seawater copper in the adjacent surf, though it was always lower (Fig. 2). The
highest eﬀective porewater copper concentration was found at Caleta Agua Hedionda (1449.59 lg Cu l1 , S.E. 842.80) which was also the location with the highest
labile seawater copper concentration (41.42 lg Cu l1 , S.E. 3.36). The lowest effective porewater copper concentration was found at Las Piscinas (6.43 lg Cu l1 ,
S.E. 0.37), whilst the lowest labile seawater copper concentration was found at
Frente Isla Pan de Azucar (1.93 lg Cu l1 , S.E. 0.24).
The densities of the meiofaunal assemblages at the reference sites (Las Piscinas,
Torres del Inca and Playa Zenteno) are clearly higher than at the other sites sampled
(Fig. 3(a)). It is also apparent that the meiofaunal assemblage densities at the
northern sites (Puerto Pan de Azucar, Frente Isla Pan de Azucar and Playa Blanca)
are higher than those from the remaining six central sites which were clearly impacted by the tailings (Fig. 3(a)). There was a signiﬁcant negative relationship between the meiofaunal assemblage densities and the eﬀective porewater copper
concentration (F ¼ 69:48, p < 0:0001) indicating clearly that meiofaunal assemblage
densities decrease with increasing copper concentrations (Fig. 3(b)).
The ANOSIM analysis for the meiofaunal assemblage densities resulted in a
global R value of 0.547 (p < 0:0001) which was statistically signiﬁcant, reﬂecting
signiﬁcant diﬀerences between sites. Post hoc pair wise comparisons were carried out
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
7
10000
-1
Copper concentration (µg Cu L )
Seawater
Porewater
1000
100
10
1
Pue
Fre
Bla
Lan
Hed
Mil
Pal
Dos
Cha
Pis
Tor
Zen
Site
Fig. 2. Eﬀective porewater copper and the labile seawater copper concentrations (lg l1 ) recorded at each
of the sites (bars represent 1SE).
by recomputing R for speciﬁc pairs of sites (Table 2). Most comparisons indicate that
the sites diﬀer signiﬁcantly from each other. The exceptions were Caleta La Lancha
with Palito 1000 m Norte and Palito 2000 m Sur, Palito 1000 m Norte with Palito
2000 m Sur, and Playa Palito with Playa Cha~
naral.
The MDS plot (Fig. 4) is the meiofaunal assemblage densities data pooled for all
sampling occasions and root transformed. This plot has a stress value of 0.08 which
indicates that it is a good representation of the similarities in the meiofaunal assemblages between the sites (see Clarke & Warwick, 1994). The degree of similarity
between two sites is represented by how close or far they are from each other on the
MDS plot. The sites on the MDS plot can be ordered as follows (from unimpacted to
impacted): Torres del Inca, Las Piscinas, Playa Zenteno, Puerto Pan de Azucar,
Frente Isla Pan de Azucar, Playa Blanca, Playa Palito, Playa Cha~
naral, Palito
2000 m Sur, Caleta La Lancha, Palito 1000 m Norte and Caleta Agua Hedionda.
SIMPER analyses were carried out using the reference, northern and central site
groupings and the within-group similarities are presented in Table 3. The four
meiofaunal groups which contributed the most to the similarity are listed. For the
northern sites, nematodes (0.197) accounted for most of the similarity of the sites
within the group, followed by harpacticoid copepods, turbellarians and foraminiferans. The average similarity between the northern sites was 0.754, and the ﬁrst four
meiofaunal groups contributed 0.619 to the similarity between these sites. Turbellarians (0.329) accounted for most of the similarity between the central sites,
8
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
Meiofaunal density 50 cm
-3
1500
1000
500
0
(a)
Pue Fre Bla Lan Hed Mil Pal Dos Cha Pis Tor Zen
Site
Meiofaunal density 50 cm
-3
3.5
3
2.5
2
y = -0.889x + 2.980
1.5
2
R = 0.874
ANOVA, F = 69.48, p < 0.0001
1
0
(b)
0.5
1
2
3
1.5
2.5
3.5
-1
Effective copper concnetration (µg Cu L )
4
Fig. 3. (a) Meiofaunal assemblage densities recorded at each of the sites using pooled data for each site
from all sampling occasion (bars represent 1SE). (b) Regression analysis of the eﬀective porewater copper
concentration and the meiofaunal assemblage densities.
followed closely by nematodes (0.328), harpacticoid copepods and foraminiferans.
The average similarity between the central sites was 0.668, and the ﬁrst four meiofaunal groups contributed 0.904 to the similarity of the central sites. For the reference sites the nematodes (0.274) accounted for most of the similarities between the
sites of the group, followed by harpacticoid copepods, turbellarians and foraminiferans. The average similarity between the reference sites was 0.746, and the ﬁrst four
meiofaunal groups contributed 0.780 to the similarity between the reference sites.
The dissimilarities between the northern, central and reference sites are presented
in Table 4, where the four meiofaunal groups which contributed most to the
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
9
Table 2
Between-site diﬀerences of 12 sites from the Cha~
naral area, for all sampling occasions, resulting from post
hoc pair wise comparisons using the analysis of similarities (ANOSIM) methodology (p-values in bold are
not signiﬁcant)
Fre
Bla
Lan
Hed
Mil
Pal
Dos
Cha
Pis
Tor
Zen
Pue
Fre
Bla
Lan
Hed
Mil
Pal
Dos
Cha
Pis
Tor
0.000
0.000
0.000
0.000
0.000
0.003
0.001
0.017
0.000
0.000
0.000
0.002
0.000
0.000
0.000
0.000
0.000
0.000
0.004
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.007
0.000
0.000
0.002
0.804
0.001
0.796
0.003
0.000
0.000
0.000
0.002
0.000
0.001
0.019
0.000
0.000
0.000
0.025
0.880
0.012
0.000
0.001
0.000
0.017
0.104
0.000
0.000
0.000
0.010
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.015
0.000
Impacted sites
Lan
Northern sites
Reference sites
Fre
Zen
Dos
Mil
Pis
Bla
Tor
Pal
Pue
Hed
Cha
Fig. 4. A non-metric multidimensional scaling (MDS) analysis of the relationship between the sites
sampled using root transformed data pooled for each site from all sampling occasions. (Stress ¼ 0.08)
dissimilarity are listed. The average dissimilarity between the northern and central
sites was 0.490. The foraminiferans contributed most to the dissimilarity (0.149)
followed by harpacticoid copepods, gastrotrichs and nematodes. The ﬁrst four
meiofaunal groups contributed 0.467 of the dissimilarity between the northern and
central sites. The turbellarians contributed only 0.057 to the dissimilarity between
the northern and central sites. The average dissimilarity between the northern and
reference sites was 0.287. The nematodes (0.113) contributed most to this dissimilarity followed by harpacticoid copepods, gastrotrichs and ostracods. The ﬁrst four
meiofaunal groups contributed 0.434 to the dissimilarity between the northern and
reference sites. The foraminiferans and turbellarians contributed 0.087 and 0.070,
10
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
Table 3
Similarities percentage (SIMPER) analysis of the contribution that the major meiofaunal groups made to
within-site group similarities (%), based on data for all sampling occasions
Sites
Meiofaunal group
%
Cumulative %
Northern Pue, Fre, Bla
Nematoda
Harpacticoida
Turbellaria
Foraminifera
0.197
0.147
0.141
0.131
0.197
0.347
0.488
0.619
Central Lan, Hed, Mil,
Pal, Dos, Cha
Turbellaria
Nematoda
Harpacticoida
Foraminifera
0.329
0.328
0.132
0.115
0.329
0.657
0.790
0.904
Reference Pis, Tor, Zen
Nematoda
Harpacticoida
Turbellaria
Foraminifera
0.274
0.232
0.161
0.113
0.274
0.506
0.677
0.780
Table 4
Similarities percentage (SIMPER) analysis of the contribution that the major meiofaunal groups made to
between-site group dissimilarities (%), based on data for all sampling occasions
Sites
Meiofaunal group
%
Cumulative %
Northern and Central
Foraminifera
Harpacticoida
Gastrotricha
Nematoda
0.149
0.122
0.113
0.084
0.149
0.270
0.383
0.467
Northern and Reference
Nematoda
Harpacticoida
Gastrotricha
Ostracoda
0.113
0.108
0.106
0.106
0.113
0.222
0.328
0.434
Central and Reference
Harpacticoida
Nematoda
Ostracoda
Foraminifera
0.212
0.174
0.116
0.108
0.212
0.386
0.502
0.610
respectively to the dissimilarity between the northern and reference sites. The average
dissimilarity between the central and reference sites was 0.488. The harpacticoid
copepods (0.212) contributed most to this dissimilarity followed by nematodes, ostracods and foraminiferans. The ﬁrst four groups contributed 0.610 to the dissimilarity between the central and reference sites. The turbellarians contributed only
0.074 to the dissimilarity between the central and reference sites.
A BIOENV analysis was conducted using the following abiotic variables: Cupw ,
Cusw , Mnsw , Nisw , Znsw Md and tailings, it was not possible to use all the combinations of variables measured due to limitations of the software (the subscripts pw
and sw refer to porewater and seawater, respectively). Only those metals which had
been identiﬁed as varying from what would be considered natural background
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
11
Table 5
Biotic/environmental variable (BIOENV) analysis showing which environmental variable, or combination
of variables, best described the biotic similarities between sites
Best variable combination (qx )
k
1
2
3
4
Cupw (0.587)
Cupw + Tailings (0.653)
Cupw + Znsw + Tailings (0.677)
Cupw + Cusw + Znsw + Tailings
(0.666)
Tailings (0.585)
Cusw + Tailings (0.642)
Cusw + Znsw + Tailings (0.662)
Cusw (0.570)
Cusw + Znsw (0.633)
Cupw + Cusw + Tailings (0.643)
k the number of variables in the combination (the best combination is in bold type).
concentrations (Lee et al., 2002) were used in this analysis. The results of the BIOENV test are presented in Table 5. The combination of variables which best explained the changes in the meiofaunal assemblages between sites was Cupw , Znsw and
tailings (0.667). The best single variable was Cupw (0.587), followed by tailings
(0.585). The best two variable combination was Cupw and tailings (0.653). Note that
the sediment grain size (Md) did not feature in any of the combinations presented in
Table 5.
The mean meiofaunal assemblage taxa diversities, expressed as the number of
groups, for all sampling occasions from January 1997 to March 2000 are presented
in Fig. 5(a). Even though the groups used are not necessarily taxonomically equivalent, the groupings are consistent throughout the study and therefore the comparison is valid. Fig. 5(a) indicates that the central sites generally had lower taxa
diversities than all the other sites. There was no distinction between the reference
sites and the northern sites in terms of taxa diversity. Regression analysis of the
relationship between taxa diversity and the eﬀective porewater copper concentration
(Fig. 5(b)) was signiﬁcant (F ¼ 33:66, p ¼ 0:0002) indicating that taxa diversity
decreases with increasing eﬀective porewater copper concentrations.
The central sites had lower foraminiferan densities than the other sites (Fig. 6(b)).
Regression analysis (Fig. 6(b)) indicates that there is a signiﬁcant decrease in foraminiferan density with increasing eﬀective porewater copper concentration
(F ¼ 17:15, p ¼ 0:002). The density of turbellarians, on the other hand, did not diﬀer
between the northern, central and reference sites (Fig. 6(c)). The regression analysis
of the relationship between turbellarian density and eﬀective porewater copper
concentration (Fig. 6(d)) was not signiﬁcant (F ¼ 3:33, p ¼ 0:098). One important
observation of the turbellarian data was that the family Otoplanidae (cf. Kata galapagoensis) often dominated the central sites, i.e. those heavily impacted by tailings.
For example, 100% of the turbellarian fauna at Caleta La Lancha and Palito 2000 m
Sur, and 99% at Palito 1000 m Norte, was represented by otoplanids.
Nematode densities at the reference sites were higher than at the rest of the sites
(Fig. 6(e)). Generally the northern sites had higher densities of nematodes than the
central sites but it was not a clear distinction. Regression analysis (Fig. 6(f)) indicated that there was a signiﬁcant decrease in nematode density with increasing effective porewater copper concentration (F ¼ 13:16, p ¼ 0:005). Finally, the
harpacticoid copepod densities (Fig. 6(g)) clearly separated the northern, central and
12
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
Meiofaunal taxa diversity 50 cm
-3
10
8
6
4
2
(a)
Pue Fre Bla Lan Hed Mil Pal Dos Cha Pis Tor Zen
Site
Meiofaunal taxa diversity 50 cm
-3
1
0.8
0.6
0.4
y = -0.288x + 0.948
0.2
R = 0.771
ANOVA, F = 33.66, p = 0.0002
2
0
0
(b)
1
2
3
0.5
1.5
2.5
-1
Effective copper concentration (µg Cu L )
3.5
Fig. 5. (a) Meiofaunal assemblage taxa diversities recorded at each of the sites using pooled data for each
site from all sampling occasion (bars represent 1SE). (b) Regression analysis of the eﬀective porewater
copper concentration and the meiofaunal assemblage taxa diversities.
reference sites. The highest harpacticoid copepod densities were found at the reference sites, and the lowest at the 100% tailings sites where they were usually absent.
There was a clear, strong signiﬁcant and negative relationship between harpacticoid
copepod density and eﬀective porewater copper concentration (F ¼ 34:90,
p ¼ 0:0002), indicating that harpacticoid copepod density decreases with increasing
copper concentration (Fig. 6(h)).
Each of the meiofaunal assemblage measures was examined in relation to the level
of tailings deposition at each site. The sites were ranked by the amount of tailings
they had received (Table 6), and the relationship between each of the measures was
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
5
(a) Foraminifera
(b)
0.6
13
y = -1.181 +0.490
4
2
R = 0.686
ANOVA,
F = 17.15, p = 0.002
0.4
3
0.2
2
1
0
0
200
2.5
(c) Turbellaria
(d)
2
150
1.5
100
Fauna density 50 cm
-3
1
50
0.5
0
0
900
3
(e) Nematoda
750
2.5
600
2
450
1.5
300
1
150
0.5
0
0
400
y = -0.447x + 1.877
2
R = 0.250
ANOVA, F = 3.33, p = 0.098
(g) Harpacticoida
(f)
y = -0.902 + 2.510
2
R = 0.568
ANOVA, F = 13.16, p = 0.005
3
(h)
2.5
300
y = -1.429 + 2.228
2
200
2
R = 0.777
ANOVA,
F = 34.90, p = 0.0002
1.5
1
100
0.5
0
Pu
e
Fr
e
Bl
a
La
n
He
d
M
il
Pa
l
Do
Chs
a
Pis
To
Ze r
n
0
Sites
0
1
2
3
4
-1
Effective copper conc. (µg Cu L )
Fig. 6. The (a) foraminiferan, (c) turbellarian, (e) nematode and (g) harpacticoid densities recorded at
each of the sites using pooled data for each site from all sampling occasion (bars represent 1SE), and
regression analyses (data Log10 ðx þ 1Þ transformed) of the eﬀective porewater copper concentration and
the (b) foraminiferan, (d) turbellarian, (f) nematode and (h) harpacticoid densities.
14
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
Table 6
The study sites ranked by a qualitative assessment of the amount of tailings present at each site
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
Table 7
Spearmans rank order correlations between the meiofaunal assemblage data and the ranked tailings impact (See Table 6 for tailings impact ranks), data collected for all sampling occasions
Measure
Correlation
Density
Taxa Diversity
Foraminifera
Turbellaria
Nematoda
Harpacticoida
)0.838
)0.810
)0.824
)0.316
)0.821
)0.952
determined using a Spearmans rank order correlation (Table 7). All the meiofaunal
assemblage measures showed strong negative correlations (< 0:800) with the tailings with the exception of turbellarian density.
4. Discussion
As only the distribution of copper appeared to be connected with the distribution
of the tailings only this metal was considered in this paper (see Lee et al., 2002). The
eﬀective porewater copper concentration and the labile seawater copper concentrations were highly correlated with each other (Lee et al., 2002), though the seawater
copper was always lower. This high degree of association suggests that the primary
source of copper at each of the impacted sites was the tailings currently deposited on
those beaches and not the ‘clear water’ tailings dumped at Caleta Palito.
It is important to understand that the between-site variation in the meiofaunal
assemblages is not as a result of the natural between-site variation in sediment grain
size (BIOENV analysis, Table 6). The sediment structure is well known as a ‘super
factor’ in determining the structure of meiofaunal assemblages (Coull, 1988). The
results presented here indicate that any eﬀect that the natural sediment grain size has
on the meiofaunal assemblages is secondary to the eﬀects of the copper mine tailings.
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
15
However, the blocking of the interstitial space of coarse sediment beaches by the ﬁne
tailings does have a signiﬁcant impact on the meiofaunal assemblages and is discussed below (see also Lee & Correa, in press).
The principal changes in the meiofaunal assemblages observed in this study are a
reduction in both density and taxa diversity with increasing porewater copper concentrations. However, the presence of tailings at a site is highly correlated with the
porewater copper concentration, as one is certainly the source of the other. It is
diﬃcult, therefore, to separate the eﬀects that copper or the tailings deposition would
have in isolation using ﬁeld data alone. There is some evidence of the eﬀects of
porewater copper alone provided by the changes to the meiofaunal assemblages at
Playa Palito, where the physical impact of the tailings is lowest amongst the impacted sites and from microcosm toxicity tests (Lee, 2001) which showed the trends
outlined above. Further evidence of the eﬀects of porewater copper alone were the
lower meiofaunal assemblage densities, but not the taxa diversities, at the northern
sites when compared with the reference sites. The eﬀects of the tailings alone were
not observable in the ﬁeld as they were always associated with increased copper
concentrations. However, microcosm tests (Lee, 2001) where varying amounts of a
tailings substitute (ﬁne sand) were added to coarse sand indicated that a reduction in
the interstitial space led to an increase in surface utilizing groups, such as foraminiferans, but a decrease in true interstitial animals, such as the polychaete Saccocirrus sonomacus (Lee & Correa, in press).
In terms of the speciﬁc meiofaunal groups the results presented here indicate that
impacted sites are characterised by the absence of the harpacticoid copepods and the
increased importance of turbellarians. The sensitivity of the harpacticoid copepods
to pollutants in general (Lampadariou, Austin, Robertson, & Vlachonis, 1997;
Sandulli & De Nicola-Giudici, 1990) and metals in particular, has been noted in
other studies (Lee, Correa, & Castilla, 2001; Van Damme, Heip, & Willems, 1984).
In most of these other studies, however, the eﬀects of the metals on harpacticoid
copepods have been confounded or masked by the presence of a mixture of pollutants. In the Cha~
naral case the absence of pollutants other than metals makes the
relationship clear in the ﬁeld.
The increased importance of turbellarians belonging to the family Otoplanidae at
the impacted sites is a new observation. This has not been reported previously in the
literature, though this may be because the focus of previous meiofauna-pollution
studies has been on the harpacticoid copepods and nematodes. Additionally, the
turbellarians are also more important members of the meiobenthos in sandy sediments, like those studied here, than in ﬁne sediments (Martens & Schockaert, 1986).
The turbellarian species encountered here was found in particularly high densities on
beaches with signiﬁcant tailings deposition. At Caleta La Lancha, for example, it
normally constituted >90% of the total meiofaunal assemblage density. The implication is that turbellarians are physiologically capable of tolerating the high levels of
labile copper encountered at these sites (Caleta La Lancha: 287.5 lg l1 eﬀective
porewater copper concentration).
A comparison of the results presented in this study with others is diﬃcult as no
other study has been conducted in a similar environment. Furthermore, other studies
16
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
of the eﬀects of pollutants in general, including metals, on meiofaunal assemblages
have been conducted in estuarine or harbour environments predominantly in Europe
(Lampadariou et al., 1997; Sandulli & De Nicola-Giudici, 1990; Somerﬁeld, Gee, &
Warwick, 1994; Van Damme et al., 1984). In these locations the physico-chemical
conditions are entirely diﬀerent to those prevailing in the high-energy sandy beaches
of northern Chile. The bioavailable component of the metals is also unknown in the
studies referred to above, as only total sediment metal concentrations were reported.
The nature of the sites, characterised by ﬁne sediments, anoxia and therefore high
levels of acid volatile sulphides, and the presence of high concentrations of organic
carbon mean that the bioavailable metals will be lower due to complexation and
adsorption, and their eﬀects possibly insigniﬁcant compared with other pollutants
present (Chapman et al., 1998). One exception is the recent study by Millward (2001)
where a detailed analysis of the porewater metal speciation was made. Their ﬁndings
with regard to both nematodes and harpacticoids are in agreement with ours. The
eﬀective porewater copper concentration range for our sites was from 6.43 lg Cu l1
(101 nM) at Las Piscinas to 1449.59 lg Cu l1 (22831 nM) at Caleta Agua Hedionda,
equivalent to the medium and high concentrations used in their experiments.
Our study demonstrates that the meiofaunal groups with the highest potential for
biomonitoring purposes in metal-enriched sandy beach ecosystems are the harpacticoid copepods and the foraminiferans. The use of harpacticoid copepods as biomonitors of metal pollution has previously been proposed by Van Damme et al.
(1984) and Lee et al. (2001). The data presented here from the Cha~
naral region of
northern Chile provides clear additional evidence that harpacticoid copepods can be
used for biomonitoring purposes. Rainbow (1997) described in detail the physiology
of metal uptake by crustaceans, indicating that for the smaller, physiologically less
advanced groups of crustaceans, including the Copepoda, the mechanisms of metal
uptake are essentially passive. It follows, therefore, that the higher the bioavailable
metal concentration in the porewater, the greater the toxic response of the harpacticoid populations. The sensitivity of harpacticoid copepods relative to nematodes in
polluted situations was proposed by Raﬀaelli and Mason (1981) in the form of the
nematode-copepod ratio for pollution monitoring. However, Lee et al. (2001)
demonstrated that the nematode-copepod ratio (when applied to high-energy sandy
beach ecosystems) was not a good measure of pollution impact, but suggested that
the mean harpacticoid density alone was a good indicator of the extent of metals
impact on sandy beaches.
In summary, we can expect that the dumping of copper mine tailings into the
coastal environment, under similar conditions to those encountered on the northern
coast of Chile, will result in a reduction in the meiofaunal assemblage density and
then taxa diversities as the bioavailable concentration of copper increases in the
porewater. Harpacticoid copepods, and possibly foraminiferans, have been highlighted as sensitive of the impact of the tailings disposal and thus have potential as
biomonitors. Finally, for the ﬁrst time the increased importance of the turbellarians
in the meiofaunal assemblage at sites impacted by metals has been recorded. The
turbellarians could therefore provide a positive biomonitor of impact to complement
the negative biomonitor of impact provided by the harpacticoid copepods.
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
17
Acknowledgements
We thank Bernardo Broitman, Evie Weiters, Marco Ramierez for assistance in
the ﬁeld. Dr Hao Zhang for the metal analyses. Thanks are also due to two anonymous reviewers who commented on an earlier version of this paper. This study was
initially supported by the International Copper Association and later by FONDAP
1501-0001 to the Center of Advanced Studies in Ecology and Biodiversity.
References
Barros, F. (2001). Ghost crabs as a tool for rapid assessment of human impacts on exposed sandy beaches.
Biological Conservation, 97, 399–404.
Boening, D. W. (1999). An evaluation of bivalves as biomonitors of heavy metals pollution in marine
waters. Environmental Monitoring and Assessment, 55, 459–470.
Castilla, J. C. (1983). Environmental impact on sandy beaches by copper mine tailings at Cha~
naral, Chile.
Marine Pollution Bulletin, 14, 459–464.
Castilla, J. C., & Correa, J. A. (1997). Copper tailings impacts in coastal ecosystems of northern Chile:
from species to community responses. In M. R. Moore, P. Imray, C. Dameron, P. Callan, A. Langley,
& S. m. Copper (Eds.), Report of an international meeting, 20–21 June 1996, Brisbane (pp. 83–92).
Australia: National Environmental Health Forum.
Chapman, P. M., Wang, F., Janssen, C., Persoone, G., & Allen, H. E. (1998). Ecotoxicology of metals in
aquatic sediments: binding and release, bioavailability, risk assessment, and remediation. Canadian
Journal of Fisheries and Aquatic Science, 55, 2221–2243.
Clarke, K. R., & Warwick, R. M. (1994). Change in marine communities: an approach to statistical analysis
and interpretation. Plymouth: Plymouth Marine Laboratories.
Coull, B. C. (1988). Ecology of the marine meiofauna. In R. P. Higgins & H. Thiel (Eds.), Introduction to
the study of meiofauna (pp. 18–38). Washington, DC: Smithsonian Institution Press.
Coull, B. C., & Chandler, G. T. (1992). Pollution and Meiofauna: Field, laboratory, and mesocosm
studies. Oceanography and Marine Biology Annual Review, 30, 191–271.
Davidson, W., & Zhang, H. (1994). In situ speciation measurements of trace components in natural waters
using thin ﬁlm gels. Nature, 367, 546–548.
Gren, I. M., Destouni, G., & Scharin, H. (2000). Cost eﬀective management of stochastic water pollution.
Environmental Modeling and Assessment, 5, 193–203.
Jaramillo, E. (1994). Patterns of species richness in sandy beaches of South America. South African Journal
of Zoology, 29, 227–234.
Jaramillo, E., McLachlan, A., & Coetzee, P. (1993). Intertidal zonation patterns of macroinfauna over a
range of exposed sandy beaches in south-central Chile. Marine Ecology Progress Series, 101, 105–118.
Lampadariou, N., Austin, M. C., Robertson, N., & Vlachonis, G. (1997). Analysis of meiobenthic
community structure in relation to pollution and disturbance in Iraklion harbour, Greece. Vie et
Milieu, 47, 9–24.
Lee, M.R. (2001). The eﬀects of the disposal of copper mine tailings on littoral meiofaunal assemblages of
the Cha~
naral area of northern Chile. PhD Thesis. University of Wales, Bangor, United Kingdom.
Lee, M. R., Correa, J. A., & Castilla, J. C. (2001). An assessment of the potential use of the nematode to
copepod ratio in the monitoring of metal pollution. The Cha~
naral case. Marine Pollution Bulletin, 42,
696–701.
Lee, M. R., Correa, J. A., & Zhang, H. (2002). Eﬀective metal concentrations in porewater and seawater
labile metal concentrations associated with copper mine tailings disposal into coastal waters of the
Atacama region of northern Chile. Marine Pollution Bulletin, 44, 956–961.
Lee, M.R. & Correa, J.A. (in press). Copper mine tailings disposal: consequences for the interstitial
polychaete Saccocirrus sonomacus (Canalipalpata, Protodrilida). Journal of the Marine Biological
Association of the United Kingdom.
18
M.R. Lee, J.A. Correa / Marine Environmental Research 59 (2005) 1–18
Martens, P. M., & Schockaert, E. R. (1986). The importance of turbellarians in the marine meiobenthos: a
review. Hydrobiologia, 132, 295–303.
McLachlan, A. (1983). Sandy beach ecology – a review. In A. McLachlan & T. Erasmus (Eds.), Sandy
beaches as ecosystems (pp. 321–380). The Hague: Dr W. Junk Publishers.
McLachlan, A., & Turner, I. (1994). The interstitial environment of sandy beaches. P.S.Z.N.I. Marine
Ecology, 15, 177–211.
Millward, R. N., Carman, K. R., Fleeger, J. W., Gamberell, R. P., Powell, R. T., & Rouse, M.-A. N.
(2001). Linking ecological impact to metal concentrations and speciation: a microcosm experiment a
using salt marsh meiofaunal community. Environmental Toxicology and Chemistry, 20, 2029–2037.
Pfannkuche, O., & Thiel, H. (1988). Sample processing. In R. P. Higgins & H. Thiel (Eds.), Introduction to
the study of meiofauna (pp. 134–145). Washington, DC: Smithsonian Institution Press.
Raﬀaelli, D. G., & Mason, C. F. (1981). Pollution monitoring with meiofauna, using the ratio of
nematodes to copepods. Marine Pollution Bulletin, 12, 158–163.
Rainbow, P. S. (1997). Ecophysiology of trace metal uptake in crustaceans. Estuarine and Coastal Shelf
Science, 44, 169–175.
Sandulli, R., & De Nicola-Giudici, M. (1990). Pollution eﬀects on the structure of meiofaunal communities
in the Bay of Naples. Marine Pollution Bulletin, 21, 144–153.
Somerﬁeld, P. J., & Clarke, K. R. (1995). Taxonomic levels, in marine community studies, revisited.
Marine Ecology Progress Series, 127, 113–119.
Somerﬁeld, P. J., Gee, J. M., & Warwick, R. M. (1994). Soft sediment meiofaunal community structure in
relation to a long-term heavy metal gradient in the Fal estuary system. Marine Ecology Progress Series,
105, 79–88.
Stark, J. S. (1998). Heavy metal pollution and macrobenthic assemblages in soft sediments in two Sydney
estuaries, Australia. Marine and Freshwater Research, 49, 533–540.
Van Damme, D., Heip, C., & Willems, K. A. (1984). Inﬂuence of pollution on the harpacticoid copepods
of two North Sea estuaries. Hydrobiologia, 112, 143–160.
Warwick, R. M. (1988a). Analysis of community attributes of the macrobenthos of Frierfjord/
Langesundfjord at taxonomic levels higher than species. Marine Ecology Progress Series, 46, 167–170.
Warwick, R. M. (1988b). The level of taxonomic discrimination required to detect pollution eﬀects on
marine benthic communities. Marine Pollution Bulletin, 19, 259–268.
Watling, H. R., & Watling, R. J. (1983). Donax serra and Bullia rhodostoma – possible bio-indicators of
trace metal pollution on sandy beaches with particular reference to the south-eastern Cape. In A.
McLachlan & T. Erasmus (Eds.), Sandy beaches as ecosystems (pp. 303–314). The Hague: Dr W. Junk
Publishers.
Zhang, H., Davidson, W., Knight, B., & McGrath, S. (1998). In situ measurements of solution
concentrations and ﬂuxes of trace metals in soils using DGT. Environmental Science and Technology,
32, 704–710.
Zhang, H., Zhao, F.-J., Sun, B., Davidson, W., & McGrath, S. P. (2001). A new method to measure
eﬀective soil solution concentration predicts copper availability to plants. Environmental Science and
Technology, 35, 2602–2607.