Aqueous speciation of Copper, Manganese, Cadmium and Zinc in the
Elizabeth River estuary (Norfolk, VA) measured using the Diffusion Gradient in
Thin-Film Technique
Michael R. Twiss1 and James W. Moffett2 [1current address: Dept. of Chemistry, Biology and Chemical Engineering, Ryerson Polytechnic University, 350 Victoria Street, Toronto, Ontario, M5B 2K3, Canada, m2twiss@ryerson.ca; 2Department of Marine Chemistry and Geochemistry, Woods Hole Oceanographic Institution, Woods Hole, Massachusetts, 02543, USA]
ABSTRACT
The Diffusion Gradient in
Thin-film (DGT) analytical technique was applied to the metal contaminated
Elizabeth River estuary, Virginia. DGT
probes were deployed in the estuary over a 6 day period, in addition to
deployment in discrete water samples collected from the same sites. Measured
DGT-labile metal concentrations, were: Cu = 4.5 to 47 nM, Mn = 104-547 nM, Cd =
284-864 pM, and Zn = 168-298 nM. Free
cupric ion concentration measured by analytical voltammetry was pCu 11.31 and
pCu 10.31 at two sites, whereas DGT estimated pCu 9.42 and 9.15,
respectively. The use of DGT appears
suitable for assessing water quality provided that the flux of organic metal
into the DGT probes can be controlled.
INTRODUCTION
Current state-of-the-art
techniques to measure the free-ion concentration of trace metal in the aquatic
environment focus on voltammetric methods, competitive ligand
exchange-adsorptive cathodic stripping voltammetry (CLE-ACSV), in particular. CLE-ACSV is a time intensive technique that
analyses discrete water samples and is sensitive to analytical interferences in
coastal waters. The Diffusion Gradient in Thin-film (DGT) developed by Davison and Zhang (Davison and 
Zhang 1994, Zhang and Davison 1995) has the potential to solve many of
the constraints that currently prevent a reliable estimate of [Mz+]
in impacted aquatic environments. The
DGT technique has the ability to provide a multi-element time-integrated
measurement of labile metal species. We
applied the DGT technique to the Elizabeth River estuary, Virginia (Fig. 1), an
area that is heavily impacted by industrial, municipal, and naval
activity. Sampling sites were selected
on the basis of an earlier study (Sunda et al. 1990) that identified a large
pollution gradient in the estuary to the near pristine conditions of Chesapeake
Bay. Our hypothesis was that the DGT
deployed in situ would provide the
same indication of water quality as the DGT used to measure the water quality
in discrete samples collected from the same site. We also tested the hypothesis that the DGT is capable of
estimating the same chemical fraction of copper as that measured by
CLE-ACSV. Our objective was to field
test the DGT technique in order to assess its ability to serve as a
cost-effective means of monitoring trace metals in the water column of
contaminated waters.
METHODS
DGT technique Figure 2 shows a schematic representation of a DGT probe. It consists of a metal chelating resin (Chelex-100, Na-form, pH 8), embedded in a hydrogel (resin gel). The resin gel is separated from the bulk test solution by a hydrogel (diffusive gel) of known thickness and small pore size (typically 2-5 nm), whose function is to control the transport of trace metals by diffusion from the test solution to the resin gel where it is fixed (Zhang and Davison 1995). A membrane filter (0.45-µm pore size) covers the diffusive gel to prevent deposition of particles on the diffusive gel that could alter its diffusive properties. The concentration of trace metal in the ambient natural water that will diffuse through the DGT hydrogel and be complexed by the resin gel is calculated as follows:
[M’] = (M · Dg) / (D · A · t) (1)
where: [M’] = concentration of DGT-labile trace metal in the bulk solution, mol·cm-3; M = mass of metal flux into the probe, mol; Dg = thickness of diffusive layer (diffusive gel plus protective membrane filter), cm; D = diffusivity of metals in aqueous solution, cm2 · s-1; A = surface area of diffusion, cm2; t = duration of deployment, s.
DGT probes used a 15% acrylamide/0.3%
bis-acrylamide diffusive hydrogel. The
diffusion coefficients used to estimate DGT-labile metal were: Cu = 5.54, Mn =
5.23, Cd = 5.42, Zn = 4.94 (x 10-6 cm2·s-1). These diffusion coefficients, determined by
the empirical flux of metal through this hydrogel were approximately 75% that
reported for seawater (Li and Gregory 1974).
Sampling Discrete depth-integrated water samples were collected at the
study sites (Fig. 1) using a peristaltic pump.
At all study sites on two dates water was collected in fluorinated HDPE
2-L bottles and returned to the laboratory.
A single DGT probe was deployed for a 6-7 h period in each of two
replicate bottles collected at each site; these deployments are referred to as in vitro deployments (Table 1A). In the
West and South Branch, DGT probes were deployed on nylon rope from buoys for 3
day periods; these deployments are referred to as in situ deployments (Table 1B). Some DGT probes served as process
controls; these were transported to the field but not deployed.
Selected water samples were
filtered (<0.2 µm) in a TeflonÒ filtration rig and stored
frozen in TeflonÒ bottles for Cu speciation
analysis by CLE-ACSV using benzoyl acetone (bzac); details of the development
of the CLE-ACSV technique using bzac in estuarine waters will be reported
elsewhere (Twiss, Moffett and Croot, in 
prep.). Total dissolved metal (Cu, Cd, Zn) was determined by anodic
stripping voltammetry following acidification and UV irradiation (1 kW, 8 hrs).
RESULTS AND DISCUSSION
The data collected allow us
to compare the level of metal contamination in various branches of the
Elizabeth River estuary that have similar salinity and pH (Table 1). The sampling design also enables a
comparison of DGT-labile metal concentrations in discrete samples from each of
the four sampling sites as well as a
comparison of measurements made on discrete depth-integrated water samples
(Table 1A) with time-integrated measurements (Table 1B) at two sites, the South
and West branches of the estuary.
The depth-integrated water
samples, into which DGT probes were deployed in the laboratory, tended to reflect a higher concentration
of DGT-labile trace metal than the DGT
probes that were deployed in the field over a continuous 6 day period (two
3-day deployments). One reason for this difference is that the depth-integrated
sample may have accumulated metal-rich
bottom waters that would have been enriched with metal fluxing from the
sediment (Skrabal et al. 1997). For the
depth-integration sampling, water was collected no nearer that 0.5 m from the
sediment, whereas the DGT probes deployed in
situ were suspended in the water column at a depth of 0.5-1.5 m from the
surface.
Table
1. Labile trace
metal in the Elizabeth River estuary during Sept./Oct. 1997 measured using the
Diffusion Gradient in Thin-film gel technique.
The sites listed correspond to study areas in Fig. 1. Values are mean ± range/2, except sites C
and D, mean ± SD (n = 3); *single value
due to DGT probes lost during deployment. A. DGT deployed in the laboratory on discrete water
samples collected in the field. B. DGT deployed in the field for 3 day periods.
|
Site |
°C |
0/00 |
pH |
Date (d/m) |
DGT-labile trace metal |
|||
|
Cu, nM |
Mn, nM |
Cd, pM |
Zn, nM |
|||||
|
A. in vitro deployments |
||||||||
|
EB |
21 |
20.8 |
7.57 |
30/09 |
9.5
± 3.4 |
547
± 113 |
864
± 61 |
168
± 40 |
|
ER |
22 |
22.0 |
7.60 |
30/09 |
14
± 5.2 |
321
± 36 |
840
± 66 |
223
± 6 |
|
SB |
23 |
20.8 |
7.52 |
30/09 |
11
± 0.5 |
393
± 50 |
578
± 40 |
243
± 41 |
|
WB |
21 |
22.4 |
7.71 |
30/09 |
11
± 3.7 |
132
± 34 |
626
± 46 |
193
± 10 |
|
EB |
19 |
21.0 |
7.62 |
03/10 |
27
± <0.1 |
506
± 3.1 |
618
± 11 |
221
± 25 |
|
ER |
21 |
22.5 |
7.56 |
03/10 |
25
± <0.1 |
212
± 5.7 |
475
± 19 |
197
± 1 |
|
SB |
21 |
21.5 |
7.54 |
03/10 |
47
± 5.0 |
351
± 25 |
665
± 35 |
298
± 29 |
|
WB |
20 |
22.0 |
7.72 |
03/10 |
22
± 4.6 |
106
± 11 |
667
± 170 |
171
± 21 |
|
B. in
situ deployments |
||||||||
|
SB |
(see above) |
30/09-03/10 |
11
± 0.6 |
273
± 9.2 |
284
± 18 |
--- |
||
|
06/10 |
21 |
21.5 |
7.54 |
03-06/10 |
10
± 0.4 |
187
± 13 |
341
± 31 |
--- |
|
WB |
(see above) |
30/09-03/10 |
4.5
± 1.4 |
113
± 1.9 |
438
± 52 |
--- |
||
|
06/10 |
21 |
22.0 |
7.97 |
03-06/10 |
5.8
± 3.1 |
104
± 1.7 |
436
± 29 |
--- |
|
SB-A |
23 |
21.5 |
7.80 |
03-06/10 |
17* |
293* |
576* |
--- |
|
SB-B |
--- |
--- |
--- |
03-06/10 |
15* |
309* |
567* |
--- |
|
SB-C |
--- |
--- |
--- |
03-06/10 |
16
± 1.1 |
298
± 27 |
578
± 43 |
--- |
|
SB-D |
--- |
--- |
--- |
03-06/10 |
18
± 2.5 |
272
± 15 |
566
± 19 |
--- |
The in vitro DGT
deployments showed the greatest variance in the DGT-labile [Cu] between sites,
compared to the in situ deployments:
e.g. SB and WB ranged from 11-47 nM and 11-22 nM, respectively (Table 1A),
whereas the same sites ranged from 11-10 nM and 4.5-5.8 nM, respectively (Table
1B), as measured by the in situ
deployments. DGT-labile Mn and Cd
showed a similar but much less marked variability between the in vitro and in situ sampling periods.
Total dissolved metal at SB (03/10) was 58 nM Cu, 180 pM Cd, and 232 nM
Zn.
In an effort to gauge the
spatial variability of water quality at a single site, 4 in situ deployments (SB-A-D) were conducted approx. 0.2 km upstream
from the SB site. At this upstream
site, deployments were made on both sides of the river within 0.1 km of each
other. Results for DGT-labile Cu, Mn
and Cd concentrations show that the water quality is relatively homogenous
within the quadrant assayed. The average coefficient of variation of replicate
DGT measurements made at a single deployment was 8%. Total Cu measured in
surface water sampled from these sites was: 38 nM SB-A; 46 nM SB-B; 44 nM SB-C;
and 58 nM SB-D.
The total dissolved copper
measured in the East and South branches match closely the measurements made in
the same season 10 years earlier (Table 2).
In addition, the Cu speciation measurements we made using CLE-ACSV are
in accordance with measurements made by Sunda et al. (1990) who employed a CLE
technique using EDTA as the competing ligand and sorption of labile copper onto
a silica coated C18 resin. However, the
DGT technique overestimated the [Cu2+] if we assume that the DGT
probes are measuring only the inorganic labile fraction of dissolved
copper. Field studies by us in other
pristine and impacted coastal seawater, in addition to laboratory study using
synthetic organic Cu complexes have revealed that the DGT technique using
Chelex is capable of removing Cu from organic Cu complexes that are small
enough to diffuse through the acrylamide hydrogel but are normally non-labile
within the time frame (<3 min) of diffusion through a 0.5 mm DGT hydrogel,
in the absence of a strong competing ligand such as Na-Chelex (Twiss and
Moffett, in prep.). Clearly, the DGT used here is not measuring only the
inorganic labile pool of copper in the natural waters.
Our study highlights the
importance of considering the flux of organic metal complexes into DGT
probes. On the basis of this field
trial, the DGT technique is considered to be well suited for assessing water
quality in a cost-effective manner, provided that the technique can be modified
so that only the inorganic labile fraction of dissolved trace metal is
detected.
Table 2. Comparative
copper speciation determined by the Diffusion Gradient in Thin-film gel
technique and Competitive Ligand Exchange-Adsorptive Cathodic Stripping
voltammetry. For DGT, [Cu2+]
was assumed to be 4% of the DGT-labile copper.
|
Site, date |
Total dissolved (<0.2 µm) Cu, nM |
pCu, -log10 [Cu2+] |
|
|
DGT |
CLE-ACSV |
||
|
EB, 30/09/97 |
29 |
9.42 |
10.31 |
|
SB, 03/10/97 |
58 |
9.15 |
11.13 |
|
Data from Sunda et al. (1990) |
CLE |
||
|
7 (EB), 30/10/87 |
28 |
--- |
11.29 |
|
8 (SB), 30/10/87 |
52 |
--- |
10.79 |
REFERENCES
Davison W, Zhang H. (1994), Nature 367: 546-548.
Li
Y-H, Gregory S (1974), Geochimica Cosmochimica Acta 38: 703-714.
Skrabal
SA, Donat JR, Burdige DJ (1997). Limnology and Oceanography 42: 992-996.
Sunda
WG, Tester PA, Huntsman SA (1990),
Estuarine, Coastal and Shelf Science 30: 207-221.
Zhang H, Davison W (1995), Analytical Chemistry 67: 3391-3400.