XUE, HANBIN. AND LAURA SIGG. Free cupric ion concentration

Transcription

Limnol. Oceanogr., 38(6), 1993, 1200-1213
0 1993, by the American
Society of Limnology
and Oceanography,
Free cupric ion concentration
eutrophic lake
Inc.
and Cu(I1) speciation in a
HanBin Xue and Laura Sigg
Institute for Water Resources and Water Pollution
8600 Diibendorf
Control
(EAWAG),
Swiss Federal Institute
of Technology,
CH-
Abstract
The free cupric ion concentrations, [Cu2+], in the water column ofeutrophic Lake Greifen were evaluated
by means of ligand exchange with catechol and cathodic stripping voltametry of the Cu-catechol complexes.
Total dissolved Cu, [Cu],, was in the range 0.5-2.8 x 10 -8 M, while [Cu2-I] ranged from lo-l6 to lo-l4
M at different times and depths. These values of [CU~-~] are much lower than literature values for the
open sea. Equilibrium models of titration data indicate the presence of very strong ligands with conditional
stability constants of 1013~y-10t4~9and corresponding concentrations of 40-90 nM.
Measured [Cu2+] was especially low during the algal bloom in spring and was higher in winter. p[Cu]
and ratios of [Cu], to [Cu2 t ] thus exhibit a seasonal pattern similar to that of algal productivity,
suggesting
that the Cu-complexing ligands are produced by algae.
Several trace metals such as Cu, Zn, and Fe
are essential to biological activity as nutrients,
but are toxic at elevated concentrations. The
biological availability of trace metals in terms
of nutrient limitation and toxicity is related to
free aquo ion concentrations and not to concentrations of total metals or metal complexes
(Sunda and Guillard 1976; Anderson and Morel 1982; Sunda and Ferguson 198 3). Trace
metals tend to be complexed by organic and
inorganic ligands in natural waters; Cu forms
stronger organic complexes than most other
divalent metals. The sensitivity of some organisms to very low concentrations of free cupric ions, [Cu2+], has been demonstrated (Sunda and Guillard 1976; Brand et al. 1986). There
is, therefore, much interest in the determination of [Cu2+] and of the concentrations and
stability constants of natural ligands. Cu complexation in natural waters has been extensively studied, mostly in seawater (Sunda and
Hanson 1987; Sunda and Huntsman
1991;
Moffett et al. 1990; Van den Berg 1984a).
Numerous techniques have been used to
measure [Cu2+] and stability constants and
concentrations of Cu-complexing ligands. Each
method has its own advantages and limitations. Voltametric
methods have been used
because they distinguish between labile species
Acknowledgments
We thank W. Sunda, W. Stumm and B. Wehrli for discussions, H. Ambiihl and collaborators of the limnology
department of EAWAG for data on algal productivity,
and
D. Kistler and A. Kuhn for sampling.
(usually inorganic species and weak organic
complexes) and nonlabile complexes (Coale
and Bruland 1988, 1990). Fixed potential amperometry avoids many of the uncertainties
related to reduction of labile organic complexes (Waite and Morel 1983; Hering et al. 1987).
Ion-selective electrodes have been used in some
cases to measure ambient [Cu2+] but are usually not sensitive enough. Bioassays (Sunda and
Ferguson 1983; Anderson et al. 1984; Hering
et al. 1987) provide direct information
on the
biological relevance of the actual concentrations, but generally lack sensitivity and represent very involved experiments. Sorption of
hydrophobic Cu complexes onto Cl8 SEP-PAK
cartridges could not completely separate complexed Cu, because only a fraction of the Cu
chelates is sufficiently hydrophobic to be extracted in this manner (Hanson et al. 1988).
Ligand-exchange techniques are based on the
competition for Cu complexation between natural ligands and added ligands, such as catechol or EDTA, and the subsequent specific determination
of concentrations
of these
complexes. [Cu2+] is then determined by equilibrium calculations. These methods allow us
to work at the natural levels of total Cu concentration, [CU],, and to determine indirectly
very low values of [Cu2+]. Ligand-exchange
techniques, combined with cathodic stripping
voltametry, CSV (Van den Berg 1984a), sorption on Cl8 SEP-PAK cartridges (Sunda and
Hanson 1987), liquid-liquid
partitioning (Moffett and Zika 1987), or chemiluminescence
(Sunda and Huntsman 199 1) have been used
1200
1201
Cu speciation in a eutrophic lake
to measure high levels of Cu complexation in
seawater. By these techniques, the presence of
strong ligands and low [Cu2+] has been demonstrated in seawater. Very little information
is available on ambient Cu complexation
in
freshwater, which often contains high levels of
dissolved organic matter (Van den Berg et al.
1987).
In this work we apply CSV to Cu-catechol
complexes (Van den Berg 1984a) to measure
very low [Cu2+] in a highly eutrophic lake. The
working conditions in these samples are carefully evaluated. Temporal and spatial variations of [Cu2+] and Cu complexation
in the
water column of a small eutrophic lake are
presented. Conditional stability constants and
concentrations of Cu-complexing
ligands are
evaluated, and the sources of strong organic
ligands that affect Cu speciation are discussed.
Theory of ligand exchange and
calculation of [Cu2+]
[CU2+]
+ [CU]i,
+
Z[CULi].
= [CU2+] + [CU],”
+ Z[Cu(cat)J
D-d. = [CU2+](1 + (Yin + ZKi[L,])
where
(2)
+ Z[CuLJ
+ Z[Cu(cat)J
+ l&m-1
+ Z ,&,JOH-1’
+ Pso4[SQ12-I.
= i,lS
(4)
where iP (ampere, A) is the peak current and
S (A M- l) expresses sensitivity. Cu-natural organic complexes do not contribute significantly to the peak current at the reduction potential
of Cu-catechol in seawater (Van den Berg
1984a) and in lake water (this study), probably
due to their stronger stability or (and) weaker
adsorption onto the electrode surface.
The free cupric ion in solution in any case
must equilibrate with all species of catechol
complexes. Therefore, [Cu2+] in the presence
of catechol can be calculated from Z[Cu(cati)]:
Z[Cu(cat)J
Icu2+l = &,Jcat2-]
+ P2cat[cat2-]2
= Wu(cat) A
(5)
CY
cat
with
a! cat
=
Pkat[cat2-l+
P2cat[cat2-12e
and
P2catrepresent the stability constants
of Cu(cat) and Cu(cat)22- complexes. The concentration
of free deprotonated
catechol,
[cat”-], can be calculated from mass balance
and the acid dissociation equilibria of catechol:
P1cat
X[Hicat] = [H,cat] + [Heat-]
= Z!l&coB[C032-]i
(3)
where Z[Cu(cat)J is the concentration of Cucatechol complexes. The catechol complexes
are determined selectively by DPCSV, together with [Cu]i,:
(1)
[Cu]i, and [CULi] represent Cu concentrations
as inorganic complexes and complexes with
natural organic ligands, respectively. Equation
1 can be rewritten as
CYin
[CU],
[CU2+] + [CU],
The determination
of [Cu2+] and complexation parameters for natural organic ligands is
based on the competition between natural organic ligands and the ligand catechol, which is
added in excess (Van den Berg 1984a). The
concentration
of Cu-catechol complexes at
equilibrium depends on the concentrations and
stability constants of the natural ligands in the
sample. After equilibrating
the sample with
catechol, the concentration
of Cu-catechol
complexes is determined by differential pulse
cathodic stripping voltametry (DPCSV); the
reduction current of Cu-catechol complexes
adsorbed onto the surface of a hanging mercury drop electrode (HMDE)
is measured.
[Cu2+] is calculated from equilibrium
relationships with the catechol complexes.
The total concentration of dissolved Cu in
the original water sample is given by
D&r =
pi represents the stability constant of the ith
Cu complex with a specified inorganic ligand
i, ai, can be calculated from the major ion
composition of lake water, Ki represents the
conditional stability constant of the Cu complex with a natural organic ligand Li, and [Li]
the concentration
of free ligand (uncomplexed); the product KJL,] is defined as the
complexing coefficient of ligand Li with Cu.
After adding catechol to a sample of natural
water and equilibrating
the ligand exchange
between catechol and natural organic ligands,
the dissolved Cu is distributed as follows:
= [cat], - {X[Cu(cat)J
+ [cat”-]
+ Z[M(cat)]}
(6)
Xue and Sigg
1202
where X[M(cat)] stands for the total concentration of catechol complexes with metal ions
other than Cu 2+. For major cations, only the
complexation constant of Mg with catechol is
available; the complexation constant with Ca
is expected to be similar. The complexes with
Mg and Ca are both negligible at pH 8. The
concentrations of catechol complexes with Cu
and other trace metals are also negligible compared to the high [cat],. Only the acid dissociation equilibria of catechol must thus be taken
into account. Equation 3 can be rewritten as
[CU],
= [CU2+](1
+ ai, + O!,at
+ I: Ki[Li]).
(7)
From Eq. 7 the sum z KJLJ can be obtained
for known [Cu], and [Cu2+] calculated according to Eq. 5
- t1 + % + &at)(8)
Concentration of free ligands and the corresponding Cu-complexing
coefficient might
remain constant after addition of catechol if
the ambient [Cu] is much lower than that of
ligands. [Cu2+] in the original sample (in the
absence of catechol) can thus be calculated from
the above complexing coefficients for equilibria between Cu and natural ligands and from
the mass balance:
[cu2+] =
D&
1 + ai, + X KJLi] ’
(9)
Experimental
Lake Greifen (Switzerland)
is highly eutrophic with a surface area of 8.5 km2 and a
volume of 150 x 1O6 m3. It has an average
depth of 17.7 m and a maximum of 32.2 m.
Water samples were collected from the deepest
point of the lake. Its tributaries are strongly
loaded with nutrients and pollutants from sewage and agriculture. Go-Flo sampling bottles
(General Oceanics, 5 liters) were used to collect
samples from different depths. Under N2 pressure, the samples were transferred to polyethylene bottles (Sigg et al. 199 1).
Sample aliquots were filtered after transport
to the laboratory. The filtration device and filtering membranes (0.45 pm) were washed with
0.0 1 M HN03 and rinsed with bidistilled wa-
ter. Aliquots of the filtered samples were immediately acidified to 0.01 M HN03. Total
dissolved Cu was measured by differential pulse
anodic stripping voltametry (DPASV). Total
dissolved Cu and Cu in unfiltered acidified
samples were also measured by graphite furnace atomic absorption spectrometry (AAS).
The filters were digested with aqua regia; Cu
in the digested solutions was measured by
graphite furnace AAS.
Dissolved organic C (DOC) was measured
by combustion at 680°C on a Shimadzu TOC
500 instrument. Samples from the lake contained 3.5 mg liter- ’ DOC. The original pH
(measured in situ with a combined sensor, Ziillig) was in the range 7.5-8.5. Total chloride
was 7 x 10m4M, sulfate 2 x 1O-4 M, alkalinity
3-4 x lOA M, and Ca l-2 x 1O-3 M (Sigg et
al. 1991).
DPCSV measurements were carried out
within 3 d after collection of samples. Water
samples used in each series of CSV measurements were stored in the dark at 4°C and filtered just before use.
A sample from lake Cristallina
(Tessin,
Switzerland), a small acidic lake (pH - 5) with
low DOC (SO.5 mg liter-‘) was measured for
comparison.
DPCSV measurements of Cu-catechol comwas performed
with an
plexes -DPCSV
HMDE, an Ag/AgCl reference electrode, and
a graphite counter electrode held in a Metrohm
VA 663 stand combined with a Metrohm E506
polarecord. Catechol was added to the samples
after 5 min of purging with Suprapure N2.
Equilibration
was allowed for 5 min under N2.
A new Hg drop was made and the stirrer
switched on simultaneously. The Cu-catechol
complexes were collected for 3 min at the electrode without applied potential. After the collection period, the stirrer was turned off and
15 s later the voltage scan was started in the
negative direction. Scanning parameters were
initial potential of 0 V (vs. the Ag/AgCl reference electrode), pulse height of 50 mV, and
scan rate of 5 mV s-l. DPCSV measurements
were carried out at 20 + 1“C. The reduction peak
potential for Lake Greifen water was - 270 + 25
mV. In the absence of catechol, no significant
current peak was measured, which means that
under the conditions used, dissolved inorganic
species of Cu do not contribute to the measured signal. Thus, the theoretical analysis of
the data could be simplified, and the first two
terms on the right-hand side of Eq. 3 were
ignored.
DPASV measurement of total Cu concentration- [Cu], was determined by DPASV in
acidified filtrates. Equipment and electrodes
were the same as those used in DPCSV. After
purging a sample with Suprapure N2 for 10
min, Cu was deposited at the Hg electrode at
a potential of -0.4 V for 3 min under stirring
and for 30 s of rest time. Anodic scanning was
performed at a rate of 2.5 mV s- l and a pulse
height of 50 mV.
Titration curve with Cu and determination
OfDPCSVsensitivity-To
obtain a Cu titration
curve, we spiked a series of subsamples with
different Cu concentrations. We pipetted 25ml subsamples into a series of 50-ml highdensity polyethylene beakers; 150 ~1 of HEPES
buffer (a solution containing
1 M N-2-hydroxyethylpiperazine-N’-2-ethanesulfonic
acid
and 0.5 M NaOH) was added to each beaker
to give a final concentration of 6 x 1O-3 M
HEPES and a final pH of 8.OkO.05. Cu was
added to all beakers but one, giving a concentration range between 7.8 x 1Oeg and 5.5 x
1O-7 M in 1O-l 5 steps. The series was allowed
to equilibrate at 20 + 1°C for 20-24 h.
DPCSV analyses were performed in the same
polyethylene beakers to minimize adsorption
effects. After purging with Suprapure N2 for 5
min, 25-50 ~1 of catechol was pipetted into the
beaker to give the desired concentration,
1 x
10m3 M in most experiments. After a 5-min
equilibration,
CSV measurement was carried
out. A titration curve was obtained in terms
of peak current (iJ as a function of [Cu], for
a single sample of lake water. The sensitivity
had to be calibrated for each individual
sample; it was determined from the portion of the
titration curve at high concentrations of Cu2+,
where stronger organic ligands were saturated,
and essentially all of the Cu was complexed by
catechol.
The titration with Cu of a blank solution
containing only KNO, (4 x 1O-3 M) and
HEPES buffer (5 x 1O-3 M), as well as catechol
(2.5 X 10B5 M), gave a linear relationship between the peak current and the Cu concentration in the range of 0.1 x 1O-g-1 x 1O-8 M
added total Cu (TCu); the blank value obtained
from this experiment was -0.05 x 1O-g M
TCu.
1203
The kinetics of ligand exchange between
added catechol and natural ligands were examined in a time-dependent
experiment.
DPCSV peak currents were followed with time
after addition of 1 X 10m3 M catechol to a
lake-water sample under N2 to protect catechol
from oxidation. The measurements yielded reproducible peak currents (&3%) over a period
of 2 h after mixing and indicated that equilibrium of the ligand exchange was reached in a
few minutes. An equilibration
time of 5 min
was thus used in all measurements.
To minimize adsorption of Cu on beaker
walls, we carried out tests of Cu-spiked lake
waters (10 pg liter-l) by DPCSV in different
beakers consisting
of different
materials,
namely Teflon, high-density polyethylene, lowdensity polyethylene, and glass. No obvious
differences in reduction current were found
among different materials at the same incubation time (20 h). The S.D. of these results
from the five beakers was +2.9%, which implies there was no serious adsorption on the
beaker walls, probably due to competition by
the strong natural ligands. High-density polyethylene beakers (50 ml) were chosen as polarographic cells (Miiller 1989).
HEPES buffer had negligible complexing effects (Good et al. 1966). No complexing effect
was detected in the titration of a blank solution
(mentioned above) as well as in the experiments with model solutions (see below). Effects
of HEPES buffer on the p[Cu] measurement
by DPCSV of catechol complexes were also
examined in the presence and absence of the
buffer at the same pH. pH was kept constant
by adding small amounts of 0.0 1 M HCl during
measurement in the absence of HEPES. NO
significant effect was found at the HEPES concentration used in the lake water. For p[Cu]
measurement in the absence of the buffer, the
pH of lake water increased up to 0.5 units over
time due to exposure to the atmosphere and
purging with N2; these pH variations would
obviously affect the results. We thus routinely
used HEPES to ensure constant pH.
p[Cu] was measured by this method in a
model solution containing 1 x 1O-8 M CU, 2
X 10B7 M EDTA, 2 x lop3 M Ca, 0.01 M
KN03, and 6 x 1O-3 M HEPES buffer, at pH
7.7. Catechol concentrations of l-5 x 1O-5 M
were used in different experiments. Measured
1204
Xue and Sigg
6
P
.-a.
43-
P
2-
1: /
0
I
-6
,
-5
I
,
-4
J
1
,
-3
I
I
-2
I
-
--t----)--t-0.4 -0.3 -0.2 -0.1 0
Potential Volt
Fig. 1. Voltametric current signals of Cu-catechol complexes as a function of catechol concentration
in terms of
DPCSV current peaks and peak current (i,) vs. total catechol concentration.
A. The current peaks (vs. potential) of
Cu-catechol in Lake Greifen water collected at 5-m depth on 8 August 1990; 1, 2, 3, 4, 5, and 6 correspond to 5 x
10e4, 1 x 10-3, 2 x 10-3, 5 x 10-3, 8 x lo-), and 1 x 10 2 M catechol without addition of Cu. B. Lake Cristallina
water collected on 7 July 1990 without addition of Cu-0;
Lake Greifen water collected at 5-m depth on 21 March
1990 without addition of Cu-O; Lake Greifen water collected at 5-m depth on 21 March 1990 with addition of 1.6
x 1O--8 M Cu--o.
p[Cu] was 12.1 +O. 1, which compares well with
the calculated value of 12.1.
The method was also applied to an algal
nutrient medium of known composition and
the results were compared with the calculated
p[Cu]. This medium contained 3.2 X 10e7 M
Cu, 3.4 x 1O-6 M EDTA, 2.4 x 1O-4 M Ca,
3.0 x 1O-4 M Mg, 7.7 x 1O-7 M Zn, 2.2 x
1O-5 M Fe, and 3.1 x 1O-5 M citrate at pH
7.6. The calculated speciation gave p[Cu] =
10.1, with most of the Cu complexed by EDTA;
p[Cu] = 9.98 was obtained experimentally
by
titration with Cu and determination of the catechol complexes at two different concentra-
tions of catechol. The measured p[Cu] was thus
in good agreement with the calculated value.
Results
Total dissolved Cu (by DPASV) was between 5 x 1O-g M and 2.8 x 10e8 M. Particulate Cu (AAS measurements) was in the range
< 0.5-2 x 1O-g M. AAS measurements of total
dissolved Cu and TCu, which are less precise
than the DPASV measurements in this range,
gave similar values. The ratio of dissolved Cu
to TCu was -0.75 to ~0.95. The distribution
coefficients Kd (ratio of solid phase concentration to dissolved phase concentration) can be
1205
Table 1. p[Cu] measured with different
1990
Lake Greifen
21 Mar
21 May
8 Aug
24 Ott
Depth
(ml
5
5
20
30
5
20
30
5
concentrations
p[Cu],
[W 1
5x10
WW
* All data wcrc obtained
14.7
15.0
14.8
14.7
15.3
15.4 .
14.4
in pH 7.5 solution
14.6
2x 10-s
5x 10-5
10.6
10.7
14.8
5
10.6
buKcred
with
HEPES.
derived from measurements in the settling particles (Sigg et al. in press) and are in the range
lo-120 x lo3 liters kg-l, comparable with
those for Lake Constance and Lake Zurich (Sigg
1987). Thus, dissolved Cu predominates in
Lake Greifen, in which the suspended matter
concentrations vary between 0.5 and 4 mg liter-l . The depth profiles of dissolved Cu during summer stratification
showed generally
higher Cu concentrations in the upper 10 m
than in deeper waters. These differences can
be explained by inputs from sewage, which
flows directly into the upper water layers.
The suitable catechol concentration must be
chosen as a tradeoff between analytical sensitivity and minimizing interferences of catechol
adsorption; 2.5 x 1O-5 M catechol was recommended for determining complexation of
Cu in seawater (Van den Berg 1984a), and 2
x 1O-4 M was used in the Scheldt estuary (Van
den Berg et al. 1987). Preliminary experiments
showed, however, that there was no detectable
current reduction for Lake Greifen water at
such low concentrations of catechol, due to
much stronger complexation
by organic ligands. Figure 1 shows the DPCSV signals in
Lake Greifen water at different catechol concentrations and the peak current as a function
of the added concentration of catechol for different water samples. The peak current increased with added catechol and approached
a maximum after the added catechol was higher than a certain concentration in water from
both Lake Greifen and Lake Cristallina. However, the critical concentration for detectable
Original
Catechol
(M)
1x10-3
15.5
14.7
15.0
14.8
14.8
15.3
15.5
14.6
15.4
22.0
12.6
13.5
9.4
23.6
8.2
7.6
13.8
7.6
by DPCSV
8X 10-d
4
1x10
Lake Cristallina*
27 Jul
of catechol (M).
pH of Lake Cristallina
6x10
10.6
5
8~10.~
10.6
water was 5.2.
current signals was dependent on the individual sample. The curve for Lake Cristallina water with organic C CO.5 mg liter-l was similar
to that for seawater. Reproducible
p[Cu]
(20.05) and good sensitivity (0.36 A M-l) were
obtained in the catechol concentration range
of 1 x 1O-5-1 x 1O-4 M for Lake Cristallina
water. The curve for Lake Greifen water (DOC
3.5 mg liter - l) was shifted toward higher concentrations of catechol by more than two orders of magnitude compared to the curve for
Lake Cristallina. The range of 5 x 1O-4-1 x
1O-3 M catechol yielded acceptable sensitivity
(0.32-0.40 A M-l) and reproducible
p[Cu]
(+O. 1 units) for Lake Greifen water (Table 1).
Therefore, a concentration of 1 x 1O-3 M catechol was chosen for determining Cu complexation in Lake Greifen water. Replications
with the same concentrations of 1 x 1O-3 M
catechol gave errors in the peak current of
~20% at 1O-lo A (ambient [Cu] in Lake Greifen water) and < 5% at 1O-8 A, corresponding
to errors of 40.1 units for pCu.
Figure 2 shows an example of the current
peaks increasing with addition of Cu and two
titrations of a Lake Greifen water with Cu.
DPCSV sensitivities were extracted from the
portion of titration curves at high [Cu2+]; the
same sensitivity was obtained for the two titrations with different concentrations of catechol (Fig. 2B). The sensitivities ranged between 0.25 and 0.40 A M-l for all samples
examined.
The titration curve with Cu (Fig. 2) gives
the concentration
of Cu-catechol complexes
1206
Xue and Sigg
200
ac
.a
100
10
20
30
40
50
60
-8
I
I
I
I
I
I
I
I
I
I
-0.4 -0.3 -0.2 -0.1 0
Potential Volt
[CuIT (M)
‘lo
Fig. 2. DPCSV current peaks (vs. potential) increasing with addition of Cu, and titration curves with Cu, in terms
of peak current (i,) as functions of [Cu], for Lake Greifen water collected at 5-m depth on 8 August 1990. A. The
current peak without addition of Cu-l;
with addition of 0.79 x lo-*, 1.26 x lo-*, 1.57 x lo-*, and 3.15 x lo-*
M Cu, in the presence of 1 x 1O-3 M catechol-2,
3, 4, and 5. B. i, measured with 5 x 1O-4 M catechol-0;
i,,
measured with 1 x 1O-3 M catechol-0.
The slopes of the lines at high concentrations
of Cu, where the stronger
organic ligands were saturated, express the sensitivities of DPCSV measurements for the corresponding water.
(from the peak current) as a function of total
added Cu; [Cu2+] was calculated for each point
from Eq. 5. The ambient [Cu2+] was calculated
with Eq. 8 and 9 from the reduction current
and total dissolved Cu in the original sample,
where ain = 57 was used, as calculated from
the major ion concentrations, with pi corrected
for an ionic strength of 0.01 M (Martell and
Smith 1974). ain is negligible compared to acal
(1 X lo’-1 X 107) or 2 Ki[Li].
Table 2 gives the [Cu2+] of the samples collected at different dates and depths in Lake
Greifen. It is surprising that [Cu”+] values are
so low compared to available values for surface
water in the open sea. The average p[Cu] at 5
m is 15.3 4 0.5, with exception of the two win-
ter samples; [Cu2+] values are, thus, two orders
of magnitude lower than most published values for surface seawater (Table 3). Only the
higher p[Cu] (12.7-14.4) for the Scheldt (Van
den Berg et al. 19 8 7) and Tamar estuaries ( 16.218.2, Van den Berg et al. 1990), which were
also determined by DPCSV, overlie the range
for Lake Greifen. The data show that [Cu2+]
values are 6-7 orders of magnitude lower than
the dissolved Cu values. Thus, almost all of
the ambient dissolved Cu in the lake is complexed with organic ligands.
Figure 3 exemplifies titration curves in terms
of [Cu2+] as a function of [Cu],. The relationship between log[Cu2+] and log[Cu], is illustrated for several sampling dates in Fig. 4. There
1207
Table 2.
(Conditional
1990/1991
[Cu],, [CL+], p[Cu], conditional
stability constants, and ligand concentrations
stability constants and ligand concentrations calculated by FITEQL program.)
Depth
(m)
[WT
(nW
of Lake Greifen
water.
[Cue’ ]
(lo-I5
M)
PKM
log K,
log Kz
(2)
21 Mar
5
22.0
0.31
15.5
23 Apr
5
20
30
4.7
15.7
17.3
0.18
0.10
0.18
15.7
16.0
15.7
14.9
12.9
40
294
21 May
5
20
30
12.6
13.5
9.4
1.88
1.03
1.61
14.7
15.0
14.8
14.1
12.3
51
254
8 Aug
5
20
30
23.6
8.2
7.6
1.52
0.53
0.29
14.8
15.3
15.5
14.3
12.1
88
539
24 Ott
5
20
30
13.8
3.2
3.2
2.32
0.66
0.47
14.6
15.2
15.3
13.9
12.2
87
556
9 Jan
5
30
27.5
13.2
8.50
18.7
14.0*
13:7*
14.1
11.8
83
396
* Obtained at pH 7.8, close to the original pH of the lake water at that time,
results determined
at pH 8 to comparc with other samples.
is an obvious shift between the two curves
represented in Fig. 4A, which indicates a difference in the complexing characteristics of the
lake water between the two samplings in April
and May; the other points at different sampling
dates also reflect changes in complexing characteristics. The two samplings in August and
October (Fig. 4B) differ in the lower part of the
curve (corresponding to the stronger ligands)
and overlap in the upper part (weaker ligands).
The direct results of these titration experiments are values of p[Cu] and of ZZKi[L,] (Eq.
2 and 9)-the product of the stability constants
and the ligand concentrations. These parameters can be estimated with reasonable precision: +O.l log units for p[Cu] and +0.2 for
log@ Ki[Li]). It is however much more difficult
to extract individual conditional stability constants and ligand concentrations from Z Ki[Li].
The complexing characteristics of these samples are certainly caused by a complex mixture
of different ligands. Approximation
of this
complex system by fitting to a simple twoligand model represents a simplification, which
allows us to evaluate the range of stability constants and of ligand concentrations that may
be present; two ligands represent the minimal
number required to fit the data.
The FITEQL program (Westall 1982), with
a two-ligand model, was used to estimate conditional stability constants and ligand concen-
.
but the constant
and ligand
concentration
were calculated
on the basis of the
trations. Most data series could be converged
with this optimization process. The computed
conditional
complexation
constants and ligand concentrations are also listed in Table 2.
500
xl O-l4
I
-
I
’
I
’
0
-
400 -
s
300 r
cu=r
200 2
100 -
0
10
20
30
[CUIT (M)
40
50
xl O-8
Fig. 3. Titration curves of Lake Greifen water with
Cu(I1) in terms of [Cuz+] as a function of [Cu],. The samples were collected at 5- (A) and 30-m (0) depths on 8
August 1990. [Cu2+] was calculated for each titration point
in the absence of catechol on the basis of ligand exchange
between the added catechol (1 x 1O-3 M) and natural
organic ligands. Insert shows an enlarged view of the lowest concentration range.
70
15
130
33
26-206
l-3
4.5
5
20
11
25-l 52
11.7-13-l
12.5-13.5
11.4
?
Sargasso Sea
8.2
8.1
Southeastern Gulf
of Mexico
Mississippi
plume
7.7
7.7
7.8
South Atlantic
Scheldt estuary
Tamar estuary
River
8.2
Coast of Peru
16.2-18.2
12.7-14.4
11.3
13.OkO.6
5+-2
2.0
12.92
8.2
Lower Newport
River estuary
North Pacific
13.2
70-80
3.3
12.54
8.1
Shelf water off
North Carolina
26
8.9
10.2
11.5-12.8
12.2
13.0-14.9
9.8
112
11.1
9.2
10.0&0.4
9.7
10.0
12.3
log Kz
stability
12.3
13.2
14.3
250-550
40-90
14.9
8.0
log K
a
Lake Greifen*
CL
P[C4
(nM), and conditional
PH
of ambient p[Cu], ligand concentrations
Sampling
location
Table 3. Comparison
Technique
and liwith
Chemiluminescence
gand competition
EDTA
comcom-
CSV of Cu-catechol
plexes
comCSV of Cu-catechol
plexes
CSV of Cu-catechol
plexes
Bioassay
Bioassay
SEP-PAWligand
competition with EDTA
DPASV
Ligand exchangeiliquidliquid partition
and liwith
com-
Chemiluminescence
gand competition
EDTA
CSV of Cu-catechol
plexes
constants.
et al. 1990
Van den Berg et al.
1990
Van den Berg et al.
1987
Van den Berg 1984b
Sunda and Ferguson
1983
Sunda and Ferguson
1983
Sunda and Hanson
1987
Coale and B&and
1988, 1990
Moffett
Sunda and Huntsman
1991
Sunda and Huntsman
1991
This study
References
&
FL!
%
ti
fz
a
1209
Cu speciation in a’eutrophic lake
-11
I
I
I
I
-11
i
B
I
I
i ’
-12
cl
- d
-15
-16
-9.0
-16
-8.0
-7.0
-6.0
-9.0
‘og [CUIT
A
-8.0
-7.0
log [CUIT
-6.0
B
Fig. 4. Titration curves of Lake Greifen water with Cu(I1) in terms of log[Cu*-+] vs. log[Cu],. The samples in panel
A were collected at 5-m depth on 23 April (0) and 21 May (0) 1990; those in panel B were collected on 8 August (0)
and 24 October (0) 1990 and 9 January 1991 (Cl). Points representing log[Cu*+] vs. log[Cu], in the original water on
the other sampling dates are plotted in panel A for comparison [21 March (A), 8 August (Cl), 24 October m, and 9
January (A)]. The curves fitted to the water data were computed from the conditional stability constants and ligand
concentrations for strong and weak organic ligands listed in Table 2.
Lo@, values for the stronger ligands range
between 13.9 and 14.9 (avg 14.3) and log Kz
for the weaker ligands between 11.8 and 12.9
(avg 12.3). These values must be considered
as estimates of the order of magnitude of the
stability constants involved, which are also influenced by the data-fitting procedure and the
assumption of a two-ligand model. The total
concentrations are -40-90 nM for the stronger ligands (CL,) and 250-550 nM for the
weaker ligands (CL,).
The titration data could be fitted well with
these two ligands, as shown by the solid curves
in Fig. 4A and B. These curves were calculated
with the stability constants and ligand concentrations given in Table 2 using the MICROQL
program (Westall 1979).
Discussion
The reliability
of the method under these
lake-water conditions must be discussed first;
the exchange kinetics with the natural organic
ligands, the effect of different catechol concentrations, and competitive effects must all be
considered.
Equilibrium
of ligand exchange between
added catechol and natural ligands is required
to determine [Cu2+] values by CSV of catechol
complexes. If equilibrium
was not attained,
Z[Cu(cat)J would be underestimated, and consequently [Cu2+] would be underestimated.
According to the kinetics of ligand exchange,
the rate of the reaction depends on the catechol
concentration; the high concentration of cat-
1210
Xue and Sigg
echo1 used should promote rapid equilibrium.
The kinetic experiment indicated that equilibrium of the ligand exchange was reached within a few minutes.
Reliable results for p[Cu] in the original water sample should be independent of the catechol concentration
used, if equilibrium
is
reached for ligand exchange between added
catechol and natural ligands. Different concentrations of catechol were thus used, but the
applicable range was limited in these samples
(Fig. 1B). For Lake Cristallina water, catechol
concentrations between 1 x 10H5 M and 1 x
10m4 M, close to that recommended by Van
den Berg (1984a) for seawater, were optimum
for determining [Cu2+]. For Lake Greifen water, the optimal catechol concentration range
was between 5 x 1O-4 and 1 x 1O-3 M, and
the measured [Cu2+] decreased at catechol
concentrations > 2 x 10m3 M. The data listed
in Table 1 show that reproducible results (_+0.1
p[Cu] units) were obtained within the applicable range. High concentrations of catechol
could introduce inaccuracies in DPCSV measurement due to adsorption problems at the
electrode surface.
The use of ligand competition
could also
introduce biases in determining
[Cu2+]. In
DPCSV technique, reaction between catechol
and metal ions other than Cu can decrease the
free catechol concentration,
resulting in low
values for [Cu2+]. Calculations for our water
conditions show that this effect was not significant.
Another problem is competition between Cu
and other metal ions in complexing with natural organic ligands. If it occurs significantly,
then the extent of Cu complexation, and therefore [Cu2+], will be affected by the free ion
concentration of the other metals. If the added
catechol also complexes these metals, the concentration of these metals bound to natural
organic ligands will decrease. The resulting increase in free natural organic ligands would
lead to underestimates of [Cu2+]. However, Cu
generally forms stronger organic complexes
than do other divalent trace metals. In addition, if such competitive effects are significant,
then the measured [Cu2+] should change with
the concentration of added catechol. Within
the given range of catechol concentration in
our experiments, the values of [Cu2+] are independent of catechol concentration (Table 1).
Therefore the above-mentioned biases were not
a problem.
Determination
of [Cu], in Lake Greifen water by DPCSV of catechol complexes appeared
to be impossible due to strong complexation
by organic ligands, unlike the situation in seawater. Van den Berg (1984b) used a high catechol concentration (8 x 1O-4 M) to overcome
interference from natural organic ligands by
determining
dissolved Cu concentration
by
standard additions. For the Lake Greifen samples, a quite high concentration of catechol (1
x 10e3 M) could not completely outcompete
the natural organic ligands for complexing Cu,
resulting in too low a measured concentration
of dissolved Cu compared to the value measured by DPASV or AAS. DPCSV measurement with a much higher concentration of catechol would have introduced interferences in
DPCSV reduction current as discussed above,
and therefore, was not feasible. Even in Lake
Cristallina water that had a much lower concentration of natural organic ligands, the dissolved Cu concentration measured by DPCSV
was lower than those obtained from DPASV
or AAS. Determination
of total dissolved Cu
by DPCSV must thus be used with caution.
An average p[Cu] (15.3) determined in samples from Lake Greifen is 2-3 orders of magnitude higher than available values for surface
water in the open sea as shown in Table 3.
This table also gives conditional stability constants and ligand concentrations for different
waters determined by different techniques for
comparison. The very low values of [CU~-~] in
Lake Greifen appear to be related to high concentrations of strong ligands. The conditional
stability constants for these strong ligands are
also higher than most values determined in
surface seawater. The stability constants in
Lake Greifen are in the same range as those
for the Scheldt estuary (Van den Berg et al.
1987), while the ligand concentrations
are
somewhat higher.
These stability constants and ligand concentrations are however, as mentioned above, dependent on the methods used, especially for
data fitting. It must be realized that natural
waters contain a wide range of different ligands
with different stability constants. The values
obtained from a simplified two-ligand model
represent an average of different ligands with
different stability constants. The results ob-
.
Cu speciation in a’eutrophic lake
tained also depend, in the case of ligand exchange methods, on the concentration of added ligand and [Cu], used in their determination
(Van den Berg et al. 1990). With the catechol
ligand exchange technique, the ligands that can
be detected depend on the value of a,,1 (Eq. 7);
a cat must be of the same order of magnitude
as x KJL,]. Using other techniques, such as
titration with Cu at higher concentration levels, may give very different results, since different ligands with a range of stability constants are present in natural waters. Depending
on the concentration range of Cu used in a
particular method, different kinds of ligands
are detected; the strongest ligands are detected
at low concentrations-close
to the natural level.
The very low concentrations of Cu2+ and
very high complexation stability constant indicate that there must be natural ligands with
strong complexation
properties in the lake.
What kinds of organic compounds can provide
such strong complexation and control Cu2+ at
such low levels?
The synthetic ligands EDTA and NTA (with
stability constants log K = 18.8 for CuEDTA
and log K = 13.0 for CuNTA) are known to
be present in Lake Greifen at concentrations
of1 x lo-*MEDTAand0.5
x 10-8MNTA
(Ulrich 1991). However, the competition of 1
x 10e3 M Ca at pH 8 results in a log conditional stability constant for CuEDTA of 11,
which is much less than those for natural organic ligands. EDTA and NTA are, thus, not
significant in complexing Cu in the lake.
Organic ligands of known structure (such as
EDTA and similar chelating ligands) have stability constants with Cu in the range of log K
= 12-l 8. The stability constants estimated in
the lake waters are however conditional for pH
8 and Ca - l-2 x 10B3 M; this means that
these ligands must have a high selectivity for
Cu over Ca. Examples of such ligands can be
found, e.g. ethylenediiminodibutanedioic
acid
(EDDS), which has a stability constant with
Cu log K = 18.4 (Mar-tell and Smith 1974); a
conditional stability constant in the presence
of Ca is calculated as log K = 15.8. Natural
ligands may achieve a similar selectivity with
chelating structures. Sulfidic groups in the organic material may contribute to the strong
complexation of Cu.
The natural organic ligands that strongly
1211
complex Cu are probably either directly or indirectly (e.g. humic materials) produced by the
biota. The occurrence of Cu chelators in algal
cultures has been demonstrated by McKnight
and Morel (1979), Van den Berg et al. (1979),
and Zhou and Wangersky (1985, 1989). An
earlier study on Cu binding by an algal exudate
from a green alga gave the conditional stability
constant and its pH dependence (Xue and Sigg
1990). On the basis of these values, the stability constant at pH 8 would be > 1013, which
is close to those we estimated for the strong
ligands in Lake Greifen water. This comparison implies that algal production of extracellular ligands is a potential source of strong Cucomplexing ligands in the lake. We have found
that proteins containing sulfidic groups were
released by a species of algae (Chlamydomonas) in response to added Cu (unpubl. work).
Analysis of organic material (Hollander 1989)
showed that organic compounds in Lake Greifen originate mainly from lacustrine algae.
Studies on the molecular weight distribution
and analytical fractionation
of dissolved organic matter in the lake (Gloor et al. 198 1;
Schneider et al. 1984) indicate that the structural and molecular weight distribution
of the
organic material varies seasonally and that
these variations parallel those of algal exudates
during algal blooms in summer.
Algal blooms in Lake Greifen occurred in
March-April
and August-October
1990, as
shown in Fig. 5 (H. Ambiihl pers. comm.),
which plots seasonal variations of assimilated
14C, chlorophyll (avg value from 0 to 5 m, data
from the Limnology Department, EAWAG),
p[Cu], and [CU], : [Cu2+] at 5-m depth. [Cu], :
[Cu2+J is equivalent to the complexing coefficient for organic ligands x KJL,] (Eq. 9) when
the terms (1 + ain) are negligible on the right
side of Eq. 9; it indicates the extent of complexation independent of [CU],. p[Cu] and
log([Cu],/[Cu2+])
exhibit the same pattern as
the variation of chlorophyll
over time. Both
have peaks in March-April,
minima in May,
and low values in fall-winter 1990-l 99 1.
A side-by-side comparison of titration curves
for log[Cu2+] vs. log[Cu], also shows stronger
complexation in the April sample compared
to the May sample (Fig. 4A), as indicated by
the shift in the curve along the x-axis (Sunda
and Huntsman 199 1). A shift between the titration curves of 8 August, 24 October, and 9
1212
Xue and Sigg
70 60 .
0
Fig. 5. Variations of chlorophyll, assimilated 14C,p[Cu],
and log([Cu],,/[Cu*+])
over time in 1990. Chlorophyll and
assimilated 14C represent averages from the values of O5-m depth; p[Cu] and log([Cu],/[Cu*+])
are the measured
values at 5-m depth. [Cu], : [Cu*-+] is equivalent to Z K,[L,]
(Eq. 9). Points do not align on the time axis due to different
sampling dates.
January is also seen (Fig. 4B); lake water sampled in January has comparatively the lowest
complexing ability. p[Cu] in January is almost
two orders of magnitude less than the peak
value in April, indicating greater variations in
p[Cu] than in [Cu],. Particulate P as a planktonic indicator also appeared to correlate with
complexation
characteristics.
Particulate
P
concentrations in the surface water in March
and April (22.6-57.4 pg liter-*) were higher
than in May (12.1 pg liter-l) or January (7.1
pugliter-‘). The above facts provide some evidence that biologically produced organic ligands play an important role in Cu complexation.
These findings in a eutrophic lake can be
compared with studies of Cu complexation in
the oceans. The typical p[Cu] levels found in
the oceans (- 12-l 3, Table 3) are often attributed to the occurrence of biologically produced
ligands (Sunda 1990; Moffett et al. 1990). Sunda (1990) suggested that an optimum level of
p[Cu] for phytoplankton
in the oceans can be
established by this mechanism, so that [Cu2+]
is neither at a toxic nor at a growth-limiting
level. In a study of Cu complexation
in the
Sargasso Sea, maximum Cu complexation was
reported in the region of the chlorophyll maximum (Moffett et al. 1990). The distribution
of Cu ligands in the North Pacific (Coale and
Bruland 1990) also suggests a biological source.
Similar mechanisms may be acting in lakes.
Our results in Lake Greifen indicate that even
lower [Cu2+] may exist in lakes than in the
oceans; the optimum level for the freshwater
biota, however, is not known. This very low
[Cu2+] in Lake Greifen appears to be related
to the very high productivity
of phytoplankton
in this eutrophic lake. Comparison with other
lakes with different trophic states and biological productivity
would show whether this is
typical.
These processes provide examples of the
profound influence that aquatic organisms can
exert on the distribution and chemistry of trace
metals in natural waters by production of extracellular organic ligands, as well as by intracellular uptake and adsorption onto cell surfaces. In turn, interactions of metal ions with
biological molecules may have important effects on the growth and physiology of organisms (Sunda 1990).
Conclusions
Our experiments indicate that ligand exchange with catechol and CSV of the Cu-catechol complexes can be used to measure ambient [Cu2+] in a eutrophic lake. The measured
values of p[Cu] at different times and depths
are in the range of - 14.0-l 6.0; [Cu2+] values
are thus - 6-7 orders of magnitude lower than
those of [CU],. Most of the dissolved CU is,
thus, complexed with organic ligands. The results indicate the presence of high concentrations of very strong ligands. We suggest that
these organic ligands are biologically produced
and that the very low [Cu2+] is thus related to
high productivity
of phytoplankton
in eutrophic Lake Greifen.
References
ANDERSON, D. M., J. S. LIVELY, AND R. F. VACCARO. 1984.
Copper complexation
during spring phytoplankton
blooms in coastal waters. J. Mar. Res. 4: 677-695.
ANDERSON, M. A., AND F. M. M. MOREL. 1982. The
influence of aqueous iron chemistry on the uptake of
iron by the coastal diatom Thalassiosira weissflogii.
Limnol. Oceanogr. 27: 789-8 13.
BRAND, L. E., W. G. SUNDA, AND R. R. L. GUILLARD.
1986. Reduction of marine phytoplankton
reproduction rates by copper and cadmium. J. Exp. Mar.
Biol. Ecol. 96: 225-250.
COALE, K. H., AND K. W. BRULAND.
1988. Copper com-
1213
tion in seawater, p. 87 l-89 1. In Trace metals in sea
plexation in the northeast Pacific. Limnol. Oceanogr.
water. NATO Conf. Ser. 4: Mar. Sci. V. 9. Plenum.
33: 1084-l 101.
-,
AND R. R. L. GUILLARD. 1976. The relationship
1990. Spatial and temporal vari-,
AND -.
between cupric ion activity and the toxicity of copper
ability in copper complexation
in the North Pacific.
to phytoplankton.
J. Mar. Res. 34: 51 l-529.
Deep-Sea Res. 37: 3 17-336.
AND A. K. HANSON. 1987. Measurement of free
GLOOR,R., H. LEIDNER, K. WUHRMANN, AND T. FLEISCH- -,
cupric ion concentration in seawater by a ligand comMANN. 198 1. Exclusion chromatography with carpetition technique involving copper sorption onto C,,
bon detection, a tool for further characterization
of
SEP-PAKcartridges.
Limnol. Oceanogr. 32: 537-55 1.
dissolved organic carbon. Water Res. 15: 457-462.
-,
AND S. A. HUNTSMAN. 199 1. The use of cheGOOD,N. E., AND OTHERS. 1966. Hydrogen ion buffers
miluminesccnce and ligand competition with EDTA
for biological research. Biochemistry 5: 467-477.
to measure copper concentration
and speciation in
HANSON, A. K., JR., C. M. SAKAMOTO-ARNOLD,D. L.
seawater. Mar. Chem. 36: 137-l 63.
HUIZENGA,AND D. R. KESTER. 1988. Copper comULRICH, M. 199 1. Modeling of chemicals in lakes-Deplexation in the Sargasso Sea and Gulf Stream warmvelopment and application of user-friendly
simulacore ring waters. Mar. Chem. 23: 18 l-203.
tion software (MASAS & CHEMSEE). Ph.D. thesis
HERING, J. G., W. G. SUNDA, R. L. FERGUSON,AND M.
9632, Swiss Fed. Inst. Technol. ETH, Zurich. 2 13 p.
M. MORE. 1987. A field comparison oftwo methods
VAN DEN BERG,C. M. G. 1984a. Determination
of the
for the determination
of copper complexation:
Baccomplexing capacity and conditional
stability conterial bioassay and fixed-potential amperometry. Mar.
stants of complexes of copper(I1) with natural organic
Chem. 20: 299-312.
HOLLANDER,D. J. 1989. Carbon and nitrogen isotopic
ligands in seawater by cathodic stripping voltammetry
of copper-catechol complex ions. Mar. Chem. 15: lcycling and organic geochemistry of eutrophic Lake
Greifen: Implications
for preservation and accumu18.
-.
19843. Determination
of copper in sea water by
lation of ancient organic carbon-rich sediments. Ph.D.
cathodic stripping voltammetry
of complexes with
thesis 89 16, Swiss Fed. Inst. Technol., ETH, Zurich.
catechol. Anal. Chim. Acta 164: 195-207.
317 p.
MCKNIGHT, D. M., AND F. M. M. MOREL. 1979. Release
-,
A. G. A. MERKS, AND E. K. DUURSMA. 1987.
Organic complexation and its control of the dissolved
of weak and strong copper-complexing
agents by algae. Limnol. Oceanogr. 24: 823-837.
concentration of copper and zinc in the Scheldt esMARTELL, A. E., AND R. M. SMITH. 1974. Critical statuary. Estuarine Coastal Shelf Sci. 24: 785-797.
bility constants. V. 1. Plenum.
-,
M. NIMMO, P. DALY, AND D. R. TURNER. 1990.
MOFFETT,J. W., AND R. G. ZIKA. 1987. Solvent extracEffects of the detection window on the determination
of organic copper speciation in estuarine waters. Anal.
tion of copper acetyl-acetonate in studies of copper(I1)
speciation in seawater. Mar. Chem. 21: 301-313.
Chim. Acta 232: 149-159.
-AND L. E. BRAND. 1990. Distribution
-,
P. T. S. WONG, AND V. K. CHAU. 1979. Meaan’d poteniial sources and sinks of copper chelators
surement of complexing material excreted from algae
in the Sargasso sea. Deep-Sea Res. 37: 27-36.
and their ability to ameliorate copper toxicity. J. Fish.
MILLER, B. 1989. Uber Adsorption von Metallionen an
Res. Bd. Can. 36: 901-905.
OberflBchen aquatischer Partikel. Ph.D. thesis 8988,
WAITE, T. D., AND F. M. M. MOREL. 1983. CharacterSwiss Fed. Inst. Technol., ETH, Zurich. 13 1 p,
ization of complexing agents in natural waters by copSCHNEIDER,J. K., R. GLOOR, W. GIGER, AND R. P.
per(II)/copper(I) amperometry. Anal. Chem. 55: 1268SCHWARZENBACH.1984. Analytical fractionation of
1274.
dissolved organic matter in water using on-line carbon
WESTALL,J. C. 1979. MICROQL1. A chemical equidetection. Water Res. 18: 15 15-l 522.
librium program in basic. EAWAG, Swiss Fed. Inst.
SIGG, L. 1987. Surface chemical aspects of the distriTechnol., Diibendorf.
bution and fate of metal ions in lakes, p. 3 19-349. In
-.
1982. FITEQL. A program for the determination
W. Stumm [ed.], Aquatic surface chemistry. Wileyof chemical equilibrium constants from experimental
Interscience.
data. Oregon State Univ.
-,
C. A. JOHNSON,AND A. KUHN. 199 1. Redox
XUE, H. B., ANDL. SIGG. 1990. Binding of Cu(I1) to algae
conditions and alkalinity generation in a seasonally
in a metal buffer. Water Res. 24: 1129-l 136.
anoxic lake (Lake Greifen). Mar. Chem. 36: 9-26.
ZHOU, X., AND P. J. WANGERSKY. 1985. Copper com-,
A. KUHN, H. XUE, E. KIEFER, AND D. KISTLER.
plexing capacity in cultures of Phaeodactylum tricorIn press. Cycles of trace elements (copper and zinc)
nutum: Diurnal changes. Mar. Chem. 17: 301-3 12.
in a eutrophic lake: Role of speciation and sedimen-,
AND -.
1989. Production of copper-comtation. In C. P. Huang et al. [eds.], Aquatic chemistry.
plexing organic ligands by the marine diatom PhaeoAdv. Chem. Ser.
dactylurn tricornutum in a cage culture turbidostat.
SUNDA,W. G. 1990. Trace metal interactions with maMar. Chem. 26: 239-259.
rine phytoplankton.
Biol. Oceanogr. 6: 41 l-442.
Submitted: 30 January 1992
-,
ANDR. L. FERGUSON.1983. Sensitivity ofnatural
Accepted: 1 December 1992
bacterial communities to additions of copper and to
cupric ion activity: A bioassay of copper complexaRevised: 29 December 1992

XUE, HANBIN. AND LAURA SIGG. Free cupric ion concentration

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