Limnol. Oceanogr., 38(6), 1993, 1227-1241
0 1993, by the American Society of Limnology and Oceanography, Inc.
Mercury biogeochemical cycling in a stratified estuary
R. P. Mason’ and W. F. Fitzgerald
Department of Marine Sciences, University of Connecticut, Avery Point, Groton 06340
J. Hurley
Wisconsin DNR, 3911 Fish Hatchery Road, Fitchburg 53711
A. K. Hanson, Jr., P. L. Donaghay, and J. M. Sieburth
Graduate School of Oceanography, University of Rhode Island, Narragansett 02882
Abstract
Total Hg in the permanently stratified Pettaquamscutt estuary was <25 pM throughout the water
column, even in highly sulfidic bottom waters. Particulate Hg was typically > 40% of the total Hg. Reactive
Hg (Hg,) was generally <3 pM and decreased with depth, but there is Hg, even in the anoxic bottom
waters. Elemental Hg (HgO) was highest in the mixed layer and below the detection limit at depth.
Demethylation is not an important source of Hg” in this estuary. Dimethylmercury was not detected.
Monomethylmercury (MMHg) was near the detection limit in the mixed layer and increased rapidly in
the low oxygen region. Dissolved MMHg correlated with bacteriochlorophyll pigments, suggesting that
the microbial community plays an important role in MMHg production in this estuary. The overall
distributions of dissolved and particulate Hg species result from the interaction with Fe and Mn redox
cycling, particulate scavenging and sinking, and MMHg production in the pycnocline. The estimated rate
of MMHg production from Hg, in the pycnocline region is 1.7% d-l. Hg” and MMHg are formed
principally in the mixed layer and in the pycnocline region, respectively. Particulate scavenging is important, and sedimentation, methylation, and HgOproduction are the principal sinks for Hg,.
Hg methylation in the water column. With improved analytical techniques (Bloom and Fitzgerald 1988; Bloom 1989) and clean sampling
procedures (Fitzgerald and Watras 1989), it is
now possible to determine the distribution of
Hg species in natural waters (Bloom and Watras 1989; Lee and Hultberg 1990). Investigations in the open ocean (Mason and Fitzgerald 1991) and in freshwater lakes (Bloom
et al. 199 1) have demonstrated the importance
of low oxygen waters in the production and
biogeochemical cycling of MMHg and dimethylmercury (DMHg). In contrast to the open
ocean, DMHg has not been detected in lakes
(Vandal et al. 1991). Mason and Fitzgerald
I Present address: Ralph M. Parsons Lab., Building 48- (1990) proposed that the labile inorganic Hg
108, MIT, Cambridge, Massachusetts 02 139.
fraction was the substrate for methylation and
elemental Hg (HgO) formation in natural waAcknowledgments
We acknowledge the assistance of Adam Cantu, Dan ters and that reactive Hg (Hg,) determination
O’Sullivan, Kathy Hardy, Jennifer Prentice, Wayne War- provided a suitable measure of this fraction.
ren, and others from the URI research group in collecting
To date, there have been no detailed invesand analyzing samples and hydrographic data, and Kristin
tigations of the distribution of Hg species in
Chaloupka for help in collecting and analyzing mercury
samples. The mercury research formed part of the Ph.D.
an estuarine system. The Pettaquamscutt esthesis of R.P.M. at the University of Connecticut. We tuary, Rhode Island, was chosen for this study
thank the reviewers for their comments.
as its physical, chemical, and biological charThe research of John Sieburth, Percy Donaghay, and Al
Hanson was supported by Rhode Island Sea Grant and by acteristics are well-known and because it conEPA Cooperative Agreement AERL 9005.
tains permanently stratified, fjordlike kettle
1227
Concern over human health risks associated
with eating fish with elevated levels of mercury
(Fitzgerald and Clarkson 199 1; Lindqvist et
al. 1991) has lead to re-examination of the
processes involved in the production and cycling of monomethylmercury (MMHg) - the
Hg compound predominating in fish. Aquatic
environments that have low inputs of Hg can
still have fish with high MMHg concentrations
(Winfrey and Rudd 1990; Lindqvist et al.
199 1). In estuaries, sulfate-reducing bacteria
are the principal methylators of Hg in sediments (e.g. Berman et al. 1990; Compeau and
Bartha 1985). Little is known, however, about
Mason et al.
’ basins with anoxic bottom waters (Fig. 1). The
basins are separated by sills 1 m deep and are
stratified as a result of vertical salinity and
temperature gradients. Vertical mixing is hindered except for aperiodic mixing due to strong
storms or unusually high spring tides. The extended oxic-anoxic transition zone (3-6 m) and
anoxic bottom waters are suitable for examining the processes that form methylated Hg
compounds. We examined whether methylated Hg species are produced under conditions
of low oxygen, and whether DMHg is formed
under estuarine conditions. As evasion of HgO
into the atmosphere is an important aspect of
the Hg cycle (Fitzgerald et al. 199 1; Vandal et
al. 199 1; Kim and Fitzgerald 1986), this study
also investigated the formation and evasion of
HgO.
r’
Narrows
Rho:
‘s,yrJ
you;”
Depth profile along channel axis of the
lower basin of the Pettaquamscutt estuary
F/,fyyJ3j
0
100
Distance
200
from
300
mouth
400
500
at Bridgetown
600
(m)
Fig. 1. Location of the sampling site and shape of the
lower basin of the Pettaquamscutt estuary, Rhode Island.
Methods
Samples were collected from a stable, permanent platform moored in the deepest part
of the Lower Pond of the Pettaquamscutt estuary (Fig. 1). Salinity, temperature, density,
pH, and Eh data from a high-resolution electronic profiler, to which the sample line was
attached, were monitored continuously allowing sampling at small intervals and providing
confirmation of the sampling depth and water
characteristics during collection (Donaghay et
al. 1992). Thirty depths were sampled between
the surface and 10 m on 4 September 1990,
and 30 depths between the surface and 16 m
Table 1. Description of the mercury species and fractions discussed.
Mercury species
Symbol
Description
His,
Determined by BrCl oxidation, followed by SnCl, reduction
Sample filtered through quartz-fiber filters before analy-
Reactive Hg
I-hit,
Dissolved gaseous Hg
Elemental Hg
Dimethylmercury
Monomcthylmercury
(dissolved)
Total particulate Hg
Particulate MMHg
DGHg
Easily reduced Hg species, usually performed on unfiltered water; represents labile ionic Hg species if corrected for any DGHg
Total volatile Hg fraction, consists of Hg” and DMHg
Determined by difference, i.e. DGHg - DMHg
Determined by GC separation of the DGHg
Determined by ethylation and GC separation
Total Hg
Total Hg (filt.)
sis
hi?’
DMHg
MMHg,,
Hg,
MM&t,
BrCl oxidation of particulate on quartz-fiber filters
Determined by alkaline digestion, ethylation, and GC
separation of particulate on quartz-fiber filters
1229
Hg in the Pettaquamscutt
BrCl
Tin chloride
reduction
oxidation
11
I
and
gold
Dilute, buffer
and ethylation
Ethylation
I
Sparge
2 ml of 30% KOH
overnight
Dichloromethane
extraction
collection
I
I
I
I
1
Sparge
and
Carbotrap
I
I
GC separation
I
Atomic
HaI
(Unf ilt.)
fluorescence
I
(Filt.
HgT
& unfilt.)
I
collection
of
I
species
I
detection
Mh
(Dissolved)
MhJ
(Particulate)
Fig. 2. Sampling protocols and analytical techniques used in the analysis of water samples for Hg species.
on 8 August 199 1. The samples were collected
in conjunction with investigations of the importance of microbes and the oxic-anoxic transition zone in the production and cycling of
radiatively important trace gases and other
trace substances in this estuary.
Hg samples were pumped directly into
rinsed, acid-cleaned Teflon bottles. The pump
operated continuously during collection. The
bottles were sealed hermetically, doublebagged, and stored in a cooler for shipment
back to the clean laboratory. The sampling period was 8 h in 1990 and 6 h in 1991. On 4
September 1990 we sampled overnight beginning at 2000 hours. The 1991 sampling began
at midday. Samples were analyzed for Hg,,
total Hg (filtered, Hg,,; unfiltered, Hg,), dissolved and particulate MMHg (MMHg, and
MMHg,), dissolved gaseous Hg (DMHg and
HgO) and, in 1991, total particulate Hg (Hg,)
(Table 1, Fig. 2). Analysis for DGHg was completed within 24 h of sampling. Two liters of
water were decanted slowly into a 2-liter bubbler and sparged for 30 min with Hg-free Ar
at 500 ml min-l. The volatile Hg compounds
were trapped either on gold columns for the
DGHg determination or on a Carbotrap-gold
train for speciation of the DGHg. The concentration was measured by cold vapor atomic
fluorescence (Bloom and Fitzgerald 1988). The
analytical methods have been described in detail elsewhere (Mason and Fitzgerald 199 1;
Vandal et al. 199 1). There was no evidence of
DMHg in any of the samples. Thus, the Hg in
the DGHg fraction was principally HgO. After
the completed analysis, the water was decanted
1230
30
Mason et al.
I
I
,
I
,
,
,
,
,
I
,
I
,
I
,
z
1
2
.I?
-0
0
2
l-
0
10
I
I
,
,
,
I
,
I
,
0
,
I .
I I
.
. I
0
0
.
,N’ 0
* #&I =30%
cb,*oo
,-’
am
I’ 0
,#’ 0 , o@
,
5 ;o
,
, %
’
.’
, I
‘. I
I
I
I
0
0
5
10
15
Total
I
20
I
25
30
(PM)
Fig. 3. Plot of the calculated Hg, vs. the measured Hg,
for the 199 1 data, illustrating precision of the analytical
methods.
back into the Teflon bottle and the bottle sealed,
double-bagged, and frozen.
Frozen samples were thawed at room temperature in the clean room. A 500-ml subsample was removed for Hg, and Hg, determination (Fig. 2). The remaining sample was
filtered through a precleaned 0.8pm quartzfiber filter held in an in-line polycarbonate
filter holder. The filter was removed and transferred to a 125-ml Teflon bottle for MMHg,
determination. In 1991, the sample was filtered through two quartz filters. One filter was
used to determine MMHg,, the other for Hg,.
The filtrate was collected in an acid-cleaned
Teflon bottle and used for Hg,, and MMHg,
measurement. Typically 0.6-l liter was filtered.
For the Hg, determination, 250 ml of the
unfiltered water was added directly to a glass
bubbler. One milliliter of a 10% acidic SnCl,
solution was added and the sample was immediately sparged to strip the reduced species
from solution and trap them on a gold column.
This procedure differs somewhat from that used
for analysis of open-ocean waters by Gill and
Fitzgerald (1987) who defined the Hg, concentration as that fraction of the Hg in an acidified sample that is reducible by SnCl, when
the analysis is performed within 24 h of acidification. In this study, the sample was not
acidified before analysis as we felt that this
could alter the distribution of Hg species in
estuarine water rich in dissolved organic and
particulate matter. The Hg, content of freshwater samples changes with time after acidification, but the analytical method adopted here
provides a good estimate of the Hg, fraction
(Bloom in press). The procedural blank was
typically 40 pg, and the detection limit, based
on three times the SD of the blank for a series
of analyses, was 0.4 pM. All Hg, analyses were
completed within 18 h of filtering the sample,
but typically within 2 h.
Oxidation with BrCl solution was used for
His-, Hgm and Hg,. One milliliter of a 0.2 M
BrCl solution was added to a 150-250-ml subsample. For Hg,, 10 ml of distilled-deionized
water (Q water) and 1 ml of BrCl were added
to the filter, contained in a Teflon bottle. After
30 min, excess oxidant was neutralized with
hydroxylamine solution before SnCl, reduction. The detection limit for the Hg, analyses
was estimated at 0.8 pM. Comparison of the
calculated total concentration (Hg,, + Hg,) and
the measured Hg, at each depth shows that
there was good agreement among most samples (Fig. 3), considering that Hg, is ~25 pM
throughout the water column (Figs. 4a, 5a).
The deviation was > 50% for three of the 26
samples and ~30% for 65% of the samples.
The agreement was generally poorer for the
anoxic water samples. Most of these samples
contained floc particles and were difficult to
subsample representatively. For these samples
the calculated total concentration is likely to
be more representative of the actual concentration than the measured Hg,, as both Hg,,
and Hg, were based on larger sample sizes.
Precision for DGHg and MMHg analyses were
similar, being between 10 and 15%.
MMHg,, was determined by derivatization,
followed by cryogenic gas chromatography
(Bloom 1989; Mason and Fitzgerald 199 1; Fig.
2). A 400-ml subsample of filtered water was
acidified and extracted by hand-shaking for 5
min with 2 x 40 ml of methylene chloride.
Extraction isolated MMHg from chloride ions
that interfere with the ethylation procedure
(Bloom 1989). The MMHg was back-extracted
into the aqueous phase, derivatized to methylethylmercury
with tetraethylborate, and
purged from solution and trapped on a Carbotrap column. The overall procedural blank,
1231
Hg in the Pettaquamscutt
Concn
0
IO
(PM)
20
30
0
Concn
(PM)
2
6
4
8
10
0
Concn
(PM)
2
6
4
Concn
8
10
0.0
(PM)
0.2
0
0.4
0
I
L-2
16
1
(4
I
I
1 2 Fluor.
10 20 sigma-t
sigma-t
F (e>
1
Fig. 4. Concentration and distribution of Hg species and hydrographic parameters (sigma-t, dissolved oxygen,
electrode potential, and fluorescence) in the lower basin of the Pettaquamscutt estuary on 4 September 1990. Symbols:
His, -0; Hgp-0; Hg,-•I; MMHg,--n; MMHg,-A.; Hg”-R.
determined by re-extraction of the water, was
typically 5 pg for a 400-ml sample. The detection limit was 50 fM. For MMHg,, 2 ml of
a 25% KOH solution was added to the 0.8~pm
quartz filter to decompose the particulate matter. After 24 h of digestion at room temperature, 100 ml of Q water followed by 2 ml of
glacial acetic acid was added to neutralize the
solution; 30-40 ml of the solution was then
Concn
0
0
10
Concn
(PM
20
30
0246810
(PM)
analyzed by the ethylation technique. This digestion method provides a quantitative recovery for fish tissue (Bloom 1989) and gave > 80%
recovery for spiked samples in this study. The
detection limit was estimated at 50 fM.
In 199 1, samples (0.3-l .2 liters) were filtered through Whatman GF/F glass-fiber filters for phytoplankton and bacterial pigment
examination. Filters were frozen and shipped
Concn
0
2
(PM)
4
Concn
6
0.0
0.2
(PM)
0
0.4
0
1 2 Fluor.
10 20 sigma-t
sig-na-t
(4
16
Fig. 5. As Fig. 4, but on 8 August 1991. (Electrode potential not determined.)
(4
1232
Mason et al.
Table 2. Recently measured concentrations of Hg in coastal waters.
Location
Rhone River
Gironde estuary and Garonne River
St. Lawrence River
Framvaren Fjord
Saanich Inlet
Concn range (PM)
Hg, 2.9-l 5.8
Hg, 22-103,
Hg ,.[: 12-38
Hg.,,l 2.4-12
Qt.1 l-8.5
Hg, 2.5-25
to the University of Wisconsin-Madison.
Samples were extracted with 90% acetone, filtered, and analyzed with reverse-phase high
performance liquid chromatography (Hurley
and Watras 199 1). Chlorophyll a (algae), pheophytin a, bacteriochlorophyll a (purple sulfur
bacteria), and bacteriochlorophyll e (brown
sulfur bacteria) were found. Water (50-l 00 ml)
was also filtered through 0.4~pm Nuclepore
filters to determine suspended particulate matter (SPM). Salinity, temperature, pH, Eh (1990
only), sigma-t, dissolved oxygen, transmission, and fluorescence were recorded before,
during, and after sample collection. Dissolved
oxygen was measured by Winkler titration.
A series of high resolution profiles along the
axis of the basin in August 1990 showed little
horizontal variability in physical, chemical, and
biological parameters along density surfaces
within and below the pycnocline. Because of
the basin structure (Fig. l), tidal influence is
restricted to the mixed layer and tidal fluctuations are typically < 20 cm (Gaines 1975).
Donaghay et al. (1992) concluded that vertical
processesdominate the water of this basin and
that horizontal advection and diffusion could
not account for the observed vertical profiles.
In addition, diel variations in CO, (Hanson
pers. comm.) and CH4 (Scranton et al. in press),
and the lack of any diel variability in salinity
and temperature demonstrate the importance
of in situ processesin shaping the vertical profiles of chemical species.
Results
The mixed layer is typically 2-3 m thick at
the sampling site (Figs. 4e, 5e). In September
1990, the density profile showed that the water
was well mixed to 2.5 m and that there was a
rapid change in density between 2.5 and 6.5
m (Fig. 4e). Below 6.5 m, the density increased
slowly with depth, as salinity increased and
temperature decreased. The mixed-layer salin:
Reference
Cossa and Martin 1991
Cossa and NoEl 1987
Cossa et al. 1988
Iverfeldt 1988
Lu et al. 1986
ity was 19.1%0,while the salinity at 10 m was
26.4o/oo.Temperature decreased from 24°C at
the surface to 9°C at 10 m. Across the pycnocline, the temperature gradient was - 3.1”C
m-l (z positive downward) and the salinity
gradient 0.73a/oom- l. The increasing salinity
and strong temperature gradient maintain a
stable stratified system (6p/6z = 1.4 kg m-4 for
the pycnocline). In August 199 1, the profiles
were similar (Fig. 5e) but with a higher density
than in September 1990, a result of both higher
temperature and salinities. The salinity increased from 19.7o/ooat the surface to 28.2o/oo
at 16 m. The density gradient was more pronounced as a result of the higher salinity and
temperature gradients (6p/6z = 2.0 kg mm4).
The Hg concentrations found in the lower
basin of the Pettaquamscutt estuary are similar
to those measured in Narragansett Bay (Hg,
11f 5 pM; Vandal and Fitzgerald unpubl. data).
Unfiltered Hg, concentrations ranged from 2
to 21 pM in 1990 and from 4 to 24 pM in
199 1 (Figs. 4a, 5a). Hg, ranged from 0.8 to 16
pM in 199 1 and from 2.2 to 13.2 pM in 1990.
These concentrations are comparable to other
recent measurements of Hg in coastal waters
(Table 2).
For Hg,, the profiles were analogous for the
two sampling periods. In 1990, Hg, concentrations decreased below 1.5 m but increased
rapidly below 3.5 m (Fig. 4a). In the anoxic
region the concentration remained elevated.
The overall profile was comparable in 199 1,
but with peaks slightly higher in the water column (Fig. 5a). The deeper waters had lower
Hg, concentrations in 199 1. The differences
are likely a reflection of storm-induced mixing
of anoxic water into the mixed layer in October
1990 (Donaghay et al. unpubl. data). There
was a partial degassing of the anoxic waters
and sulfur precipitation in the surface waters.
As a result, CH, concentrations were lower in
August 199 1 than in September 1990 (Scran-
.
Hg in the Pettaquamscutt
1233
ton et al. in press). Gaines (1975) showed that layer to 3-5 nM in the anoxic bottom waters,
the sulfide concentration of the bottom waters while organic Cu concentrations, determined
by solid-phase separation techniques, dehad not returned to prestorm concentrations
creased from 3-5 nM in the mixed layer to
(4 mM) 15 months after a period of mixing
< 15 pM at depth, even though DOC concenand overturn in November 197 1. Immediately
after mixing, the sulfide concentration was 0.3 trations were similar throughout the water colmM-an order of magnitude lower than the umn. It is likely that the Cu, which is reduced
prestorm concentration (Gaines 1975). It is in sulfidic waters (Dyrssen and Kremling 1990),
therefore likely that the concentration of Hg is bound by sulfides, rather than DOC, in the
species in the bottom waters had not returned anoxic waters. Particulate Cu (Cu,), absent in
the mixed layer, increased to a maximum of
to the prestorm values by August 199 1.
Hg, decreased from 40% of Hg, at the sur- 9 nM in the anoxic waters. In the anoxic
face to < 10% at 2.5 m (1991 data; Fig. 5a). regions, particulate Cu, was > 50% of the total
Cu,, a percentage similar to that found for
Below 2.5 m, Hg, constituted a large fraction
of the Hg, (5 l&2 1%; n = 23). The results for Hg,. Thus, similar processes- organic com1990, based on the Hg, and Hg,, concentra- plexation in the mixed layer, sulfide complextions, were similar with Hg, being the major ation in anoxic waters-control the speciation
fraction below 3.85 m (Fig. 4a). MMHg, ranged of Cu and Hg in this estuary.
Hg, was generally below 3 pM, with the
from the detection limit (50 fM) in the mixed
layer and anoxic waters to 2.92 pM in the pyc- highest concentrations occurring in the mixed
nocline in 1990 and to 6.88 pM in 1991. The layer. Concentrations decreased overall with
depth of maximum concentration was similar depth (Figs. 4b, 5b); in the deeper waters they
were near the detection limit of 0.4 pM. For
for both sampling periods (Figs. 4c, 5~).
The difference between Hg, and Hg, cannot Hg,, the general decrease in concentration in
be attributed solely to Hg, which, on average, the anoxic region with increasing sulfide folaccounted for -30% of the Hg,, (Figs. 4, 5). lows thermodynamic predictions (Dyrssen
1989) and is likely a result of complexation of
Estimated “dissolved strongly complexed Hg”
(i.e. HgT, - Hg, - MMHg,) was highest in Hg by sulfide ligands and subsequent scavengthe mixed layer (6.6k3.9 pM for O-3 m, n = ing by FeS precipitation (Dyrssen and Krem7; 1991 data) and was relatively constant in ling 1990). Trace metals, scavenged from the
the deeper waters (2.6k2.0 pM, n = 2 1). The oxic waters by Fe oxyhydroxide precipitation,
are remineralized with Fe in the low-oxygen,
0.8~pm quartz-fiber filters used for filtration
do not collect colloidal matter and thus a frac- low-sulfide region. At high sulfide, however,
tion of the dissolved strongly complexed Hg FeS precipitates and trace metal-sulfide comis likely to be Hg associated with colloidal ma- plexes are simultaneously scavenged from soterial. There is no relationship, however, be- lution, maintaining low trace metal concentween Hg, and Hg, (r = 0.19), suggesting that trations in sulfidic waters (Dyrssen and
Kremling 1990).
higher Hg,, and presumably higher colloidal
matter (Honeyman and Santschi 1988), does
The Hg, data indicate that there are labile
not necessarily coincide with higher Hg,. Thus, Hg species even in the highly sulfidic waters
colloidally bound Hg is not a predominant
of the estuary. Thermodynamic calculations
fraction of the Hg,,. Strong complexes with indicate that the measured concentration of
dissolved organic matter, derived principally
dissolved Hg is below the intrinsic solubility
from freshwater input (log K = 18-20 for Hg- of Hg in the presence of HgS solid (Dyrssen
humic acid associations, Mantoura et al. 1978), 1989) and that all the ionic Hg is bound as
also contribute to the dissolved Hg pool in the sulfide complexes. Some of these sulfide speupper reaches of the water column. In addi- cies must be “reactive” (i.e. reducible by SnCl,).
tion, dissolved Hg in the anoxic waters is likely The Hg, determination used here is a suitable
to be bound as sulfide species [log K > 30 for measure of the Hg substrate available for
HgS, Hg(SH),, HgSZ2-; Dyrssen 19891.
methylation, HgO formation, and other conA study of Cu in the upper basin of the es- version processes(Mason and Fitzgerald 1990).
tuary (Mills et al. 1989) found that total dis- Therefore, there is available Hg substrate even
solved Cu decreased from - 8 nM in the mixed in anoxic waters of the estuary. The presence
1234
Mason et al.
of available Hg substrate could explain why
Hg methylation occurs, for example, after adding HgS to slightly acidic sediments (pH 6;
Fagerstrom and Jernelijv 197 1) and suggests
that the notion that Hg is bound as HgS and
unavailable in anoxic waters is not correct.
Dissolved MMHg was near the detection
limit (50 fM) in the mixed layer (upper 2 m),
but increased as oxygen and Eh decreased, to
a maximum in the pycnocline (Figs. 4c, 5~).
Concentrations remained elevated in the anoxic waters. Similar concentrations have been
found in the northern Wisconsin (Bloom and
Watras 1989) and Swedish lakes (Lee and
Hultberg 1990). In 199 1, the MMHgD concentration of the deeper waters generally decreased from -0.9 pM (5.5-6.5 m) to ~0.2
pM between 14 and 16 m. Although no samples were collected in the deeper waters at this
20-m-deep site due to the limitations of the
CTD-pumping system, the results indicate little diffusion of MMHgP from the bottom sediments into the anoxic zone. Rather, they suggest that there is diffusion of MMHgD from the
pycnocline region into deeper waters. The similarity between the distribution of Hg species
in stratified lakes and the Pettaquamscutt estuary indicate that similar processes control
the formation and cycling of Hg in these aquatic systems.
the presence of the phototrophic bacterial
community and that these bacteria could be
associated with Hg methylation in the pycnocline of the Pettaquamscutt estuary.
Although there is no evidence in the literature of sulfide-oxidizing bacteria methylating
Hg, methylation under aerobic conditions has
been demonstrated, both in marine (Topping
and Davies 198 1) and freshwater environments (Winfrey and Rudd 1990), most likely
by aerobic microorganisms. However, as pigmented sulfur bacteria oxidize sulfide, their
photosynthetic metabolism is intimately tied
to the use of reduced sulfur compounds, and
thus they are often found associated with dissimilatory sulfate-reducing bacteria (Madigan
1988) at the redox interface. These sulfate-reducing bacteria, which have been shown to be
the principal Hg methylators in estuarine sediments under laboratory conditions (Compeau
and Bartha 1985), could be forming MMHg
in the pycnocline region. The deeper maximum of MMHgr, at 5 m, both in 1990 and
199 1, below the depth of penetration of photosynthetically active radiation and in the region of increasing sulfide (Donaghay and Hanson unpubl. data) could reflect methylation of
Hg by these bacteria.
The maximum of MMHg, occurred 0.1-0.3
m higher in the water column than the first
MMHgD peak and coincided with the rapid
Sources and sinks for methylmercury
change in Eh (e.g. 1990 data; Fig. 4e). In 1990,
There are two probable sources of the high the peak in MMHg, was at 4.4 m with undepycnocline concentrations of MMHg: either tectable concentrations at 4.3 m. In 199 1,
MMHg is produced in situ at this depth or the MMHg, peaked just above the peak in overall
distribution of MMHg is controlled by redox fluorescence and bacteriochlorophylls (Fig. 6).
chemistry and precipitation-dissolution
is The MMHg, profile indicates that MMHg, is
formed under rather specific conditions that
controlling the distribution at the oxic-anoxic
interface, as is found for Fe and Mn. There is exist over a small depth interval within the
a significant correlation, for the 199 1 data (n redox interface. Particulate MMHg could be
= 27), between MMHgD concentration and associated with bacteria if MMHg is produced
both bacteriochlorophyll a (BChl a, r = 0.92) by these organisms.
The region of rapid change in bacteriochloand bacteriochlorophyll e (BChl e, r = 0.85),
and no correlation between MMHgD and Chl rophyll pigments is 3.5 5-4.5 5 m, similar to
a (r = 0.15) or pheophytin (r = 0.37) (Fig. 6). that of MMHg,, and there is a weak correlation
The principal maxima in MMHgD and in bac- between these parameters (r = 0.70 for BChl
teriochlorophyll occur at the same depth (4.5 5 a; r = 0.44 for BChl e). The highest Hg, conm), and there is little MMHg, in the mixed centrations also coincide with the highest baclayer where bacteriochlorophyll
concentra- teriochlorophyll concentrations (Figs. 5a, 6a).
tions are below the detection limit. In the deep- MMHg, was < 10% of the Hg, in the mixed
er waters (> 6 m), bacteriochlorophyll decreas- layer (< 2 m) and in the deeper waters (> 6 m),
es with depth, similar to the MMHg, profile. but was the major part of Hg, at other depths.
These results suggest that MMHgD forms in The highest %MMHg/Hg, (9 1%) occurred at
1235
Hg in the Pettaquamscutt
Concn
0
2
4
Relotive
(PM)
6
a
10
0
10
Concn
concn
20
30
CUM)
40
t
1990 Data
1
Fig. 6. Distribution of MMHg,, (A) and MMHg, (A), pigments (Chl a --O; BChl a-m; BChl e-A),
and Mn,, (0) on 8 August 1991.
and Fe,, (0)
4.25 m, coincident with the maximum
able for the Pettaquamscutt, but dissolved Fe
MMHg, but slightly higher than the bacterio- and Mn have been measured on the 1990 samchlorophyll maxima at 4.55 m and the Hg, ples (Fig. 6). The concentration of dissolved
maximum at 4.7 m. Bloom et al. (1991) found Fe and Mn increases rapidly below 3.9 m,
a similar peak in %MMHg,/Hg, at the redox reaching a maximum around 5 m. The peak
interface (8 m) in Little Rock Lake, Wisconsin. in Fe, should occur in the low-oxygen, lowHurley et al. (199 1) showed that the highest sulfide region, above the peak in Fe, (i.e. -4
concentrations of particulate and dissolved Hg m). FeD precipitates as it diffuses up out of the
in this lake during summer stratification oc- anoxic region into the low-oxygen upper pycnocline waters and it is possible that the precurred at the oxic-anoxic interface, coincident
with the maximum for chlorophyllous pig- cipitation of Fe is scavenging Hg and MMHg
ments and total unfiltered Fe. Below the redox from solution, resulting in the MMHg, peak
interface, MMHg, concentrations decreased higher in the water column than the MMHgD
while Hg, concentrations remained elevated, peak. The rapid change in MMHg, concensimilar to the Hg, speciation in the Petta- tration in the zone of high MMHg, and the
quamscutt anoxic zone. It is likely that MMHg,
relatively low mixed-layer MMHgD suggestthat
is formed where the relative concentration is MMHg is being removed by particulate scavmaximal, in the region of rapid change in ox- enging and sinking at this depth.
ygen and Eh.
There should also be a peak in Hg,, in this
It is possible that Fe and Mn cycling at the region. There is an increase in Hg, concentraredox boundary is influencing the distribution
tion from 2.0 pM at 3.55 m to 9.2 pM at 4.0
of Hg and MMHg. Erel and Morgan (1991) m, and Hg, concentrations remain elevated
demonstrated the importance of adsorptionthroughout the pycnocline zone. The profile
desorption from particle surfacesfor trace metal for Hg, is less readily interpreted than that of
cycling, and showed that, in seawater, Hg has MMHg, as there is a significant flux of Hg,
a strong affinity for oxide surfaces. Haraldsson from the mixed layer into the pycnocline (flux
and Westerlund (1988) suggested that Fe cy- > 125 pmol rnw2 d-l for 5 pM Hg,; sinking
cling influences the distribution of Cu, Cd, Pb, particulate-suspended matter = 0.025 and
Co, and Zn in Framvaren Fjord and the Black sinking at > 1 m d- l). Hg, is likely being formed
Sea. These data support the notion of Fe con- by Fe and Mn precipitation, and Hg, will also
trol over Hg cycling at the redox interface. At be associated with living organisms present in
present, no particulate Fe or Mn data are avail- this region. Overall, the profiles for the Hg,
1236
Mason et al.
species (MMHg, and Hg,) are consistent with
the premise of scavenging of Hg and MMHgD
by Fe and Mn precipitation.
In addition, the lower MMHg, peak (at 4.95
m in 1990) corresponds to the broader maximum in dissolved Fe and Mn (4.75-5.3 m;
Fig. 6) suggesting that MMHgr, is released as
particulate matter dissolves in the anoxic
regions in a similar manner to Fe. There should
also be a peak in Hg, in this region if Fe and
Mn dissolution is releasing significant amounts
of Hg into solution. No distinct peak was found
(Figs. 4b and 5b). However, complexation of
Hg by dissolved organic complexes could mask
any increase due to particulate dissolution.
There is an increase in “dissolved nonreactive
Hg” (I.e. Hg,, - H&t - MMHg,) slightly
deeper in the water (between 5.5 and 6 m)
indicating that Hg, is released into solution at
deeper depths than MMHg,. In addition, Hg,,
remains elevated in the deeper anoxic waters,
suggestive of strong complexation of Hg by
sulfide ligands, similarly for MMHg (log K =
2 1 for CH,HgS; Dyrssen 1989). The relatively
high MMHg, concentrations in this region
could reflect strong complexation by sulfide or,
alternatively, MMHg, production throughout
the anoxic zone, or in the sediments, by sulfate-reducing bacteria.
These observations and correlations provide
a consistent explanation for the overall distributions of MMHg, and MMHg,. MMHg is
formed in situ in the presence of the bacterial
community in the lower pycnocline region, resulting in the elevated concentrations of MMHg
in this zone. Although this investigation has
provided further evidence that MMHg is
formed in the water column by the bacteria
associated with sulfur cycling at the redox interface, it has not identified the microbial community responsible for MMHg production.
Even though the internal cycling and distributions of MMHg in the pycnocline are complex, it is possible to estimate the net rate of
formation of MMHg required to balance the
losses of MMHg from this region (2.5-6 m).
The MMHg, concentrations are low both in
the mixed layer and anoxic waters (199 1 data),
and thus the flux of MMHg, into the pycnocline region is likely to be similar to the flux
out, assuming a constant sinking velocity for
particulate matter. In addition, the concentration of MMHg, is low in the mixed layer and
there is little gradient in MMHg, concentration above - 3.5 m. Thus diffusion of MMHgD
into the mixed layer from the pycnocline region is not a significant flux of MMHg,, as
MMHg, is scavenged by precipitation of Fe
and Mn in the upper reaches of the pycnocline
(-4 m). The predominant diffusive flux of
MMHg, is out of the thermocline region into
the deeper anoxic waters. The flux is estimated
at 100 pmol m-2 d- I, assuming a diffusion
coefficient of 5 X lo+ m2 s-l and the concentration change between 5.5 and 8 m (240
pmol m-4; Fig. 4~). The diffusion coefficient
tias estimated with the relationship between
the buoyancy gradients [W = g( l/p)@p/8z) =
10e2 sm2here] and diffusion coefficient (Sarmiento et al. 1976).
Demethylation is not an important sink for
MMHg below the pycnocline as the Hg” concentration decreases to the detection limit below 4 m, suggesting there is little Hg” formation and consequently little demethylation.
Thus the diffusive flux of MMHg into the anoxic waters is the predominant flux of MMHg
from the pycnocline region and can be used as
an estimate of the overall net formation of
MMHg in this zone. The net MMHg formation
rate (expressed as the rate of conversion of the
available substrate) is estimated to be 1.7% d- l
(2 X 10- 7s- I), assuming formation occurs over
the whole region (3-m-thick pycnocline) and
assuming an average pycnocline Hg, concentration of 2 pM.
This rate is substantially higher than the specific methylation rate (i.e. the amount of added
Hg converted) measured in laboratory culture
experiments with lake water and high concentrations of Hg (O.Ol-0.3% d-l; Gilmour and
Henry 199 1) and is also higher than the estimated formation rate for methylated Hg species in the low-oxygen region of the equatorial
Pacific Ocean (0.01% d-l; Mason 1991). Although the estimated methylation rate in the
Pettaquamscutt estuary (1.7% d- ‘) is higher
than the rates estimated in the laboratory, the
actual rate of formation of MMHg (-0.07 ng
liter-’ d-l) is considerably slower than that
measured during the laboratory exposure experiments (- 1 ng liter-’ d-l; Gilmour and
Henry 199 1). Thus, water-column methylation could easily sustain the relatively high
specific methylation rate of 1.7% d- I, even if
actual methylation rates in the Pettaquamscutt
Hg in the Pettaquamscutt
basin are lower than those measured in laboratory experiments.
Air-water exchange of mercury
DGHg measurements were made in 1990
and 1991. There was no evidence of DMHg
in these waters, and therefore the DGHg consists primarily of HgO. Mason and Fitzgerald
(1990) reported evidence of DMHg in the lowoxygen waters of the equatorial Pacific, but
there have been no reports of DMHg based on
positive identification in freshwater systems
(Vandal et al. 199 1; Bloom pers. comm.). Laboratory experiments indicate that the stability
of DMHg decreases with increasing temperature and decreasing pH (Mason 199 1) but that
DMHg is still unstable even in the low-oxygen,
cold, high-pH environment of the subthermocline equatorial Pacific Ocean. Thus, the
lack of DMHg in the Pettaquamscutt estuary
suggeststhat, if DMHg is formed, the rates of
formation and decomposition must be similar.
At present, there is insufficient evidence to determine whether DMHg is formed in all aquatic systems and to assesswhich conditions enhance the stability of DMHg so that it can
accumulate to detectable levels.
In 1990, concentrations of Hg” were at the
detection limit (~25 fM) for the mixed layer,
while concentrations increased in the pycnocline region (Fig. 4d). The lack of Hg” in the
mixed layer can be attributed to a degassing
of the surface waters due to enhanced mixing
of the upper layers during a strong northeasterly storm the day before measurements were
made (3 September 1990). The wind, blowing
along the axis of the river, deepened the mixed
layer and caused the intrusion of oxygenated
water into the low-oxygen pycnocline region
(Donaghay et al. unpubl. data) as indicated by
the in situ fluorescence and Eh spectrum obtained (Fig. 4e). The bimodal peak in fluorescence in 1990 is not evident in the August 199 1
profile (Fig. 5e) which shows a distinct maximum in fluorescence at -4.5 m. In addition,
the area of elevated fluorescence is somewhat
broader in 1990, but the intensity of the maximum is greater in 199 1.
If excessive mixing of the water column
caused a degassingof all the Hg” from the mixed
layer, it would take at least 1 d before Hg”
concentrations built up to detectable levels (25
fM). This estimation was made with an av-
1237
erage Hg, concentration of 3 pM, a conversion
rate of 5% d- 1 (Mason 199 1; Vandal et al.
199 l), and an assumed negligible loss of Hg”
from the surface waters. The presence of Hg”
below 2.5 m and the similarity between the
1990 and 199 1 profiles below 2.5 m (Figs. 4d,
5d) indicate that mixing caused little loss of
Hg” from below the mixed layer. Concentrations decreased rapidly in the anoxic region
during both years.
There was measurable Hg” in the mixed layer in 1991 (190&45 fM, n = 5) and the concentration was uniform in the upper 2 m. The
flux of Hg” upward from the subsurface maximum at 2.8 m is estimated at 130 pmol m2
d-l with a d’ff
i usion coefficient of 5 X lop6 m2
s-l (W = 0.8 x 1O-2 ss2). The flux to depth
is similar (100 pmol m-2 d-l). The flux at the
surface due to gas exchange is estimated at
100-200 pmol m2 d-l, assuming a gas exchange coefficient of 0.5-1.0 m d-l (i.e. for
windspeeds of 12-l 9 km h-l). This calculation
suggeststhat if the system was at or near steady
state during the 199 1 sampling and gas evasion
is the only sink for Hg” in the estuary, then
there is little formation of Hg” within the upper
2 m of this basin. In addition, the shape of the
profile in 199 1 and the elevated concentrations
in the pycnocline region relative to the deeper
waters in 1990 and 199 1 indicate that Hg” is
formed primarily at the bottom of the mixed
layer and in the upper reaches of the pycnocline. This finding suggeststhat Hg” formation
is not a surface layer, photochemically induced
reaction.
In 199 1, the production rate necessary to
maintain the Hg” profile can be estimated at
230 pmol m-2 d-l or 3% d- 1 for a 2 pM Hg,
concentration and assuming production in the
upper 4 m of the water column, If Hg” production is restricted to the 2-4-m region, as
suggested above, the formation rate would be
double this estimate. Examination of the fluxes
of MMHg showed that demethylation is unlikely to be an important source of Hg”
throughout the water column, and thus Hg” is
formed principally by direct reduction of ionic
Hg. There is some correlation between Hg”
concentration and Chl a (r = 0.6, n = 16) suggestive of a phytoplankton role in Hg” production. However, this correlation could also
be a reflection of the fact that concentrations
of both species are highest in the mixed layer
1238
Mason et al.
Atmospheric
Gas
evasion
T
150
deposItIon
I
170
130
Wet
Dry
Surface
ProductIon
of Hg”
Productlon
of
MMHg
All
fluxes
in
pmol m2d-’
Fig. 7. Model of Hg cycling in the lower basin of the
Pettaquamscutt estuary.
and upper pycnocline and low in the anoxic
waters.
Studies in the Wisconsin lakes have also
shown that Hg” concentrations are generally
lower in the anoxic regions compared to oxic
waters (Vandal et al. 1991), as found in the
Pettaquamscutt estuary. In addition, in lakes
with significant thermocline productivity in
summer (and a resultant middepth oxygen
maximum), the maximum Hg” concentration
was at middepth, suggestive of a biological role
in Hg” production (Vandal et al. 199 1; Vandal
unpubl. data). The average Hg” % saturation
(-600%) for the mixed layer of the Pettaquamscutt is similar to the lower range of %
saturation found in the Wisconsin lakes in
summer (Vandal et al. 1991) and in the equatorial Pacific Ocean (Mason 199 1). The differences in the fluxes from the various bodies
of water depend strongly on the wind regime,
with highest fluxes estimated for the equatorial
Pacific where winds are typically stronger. The
estimated flux from the Pettaquamscutt estuary (100-200 pmol me2 d- ‘) is similar to fluxes
from the lakes in Wisconsin (Mason 199 1).
Mercury cycling in the Pettaquamscutt estuary
Atmospheric wet deposition would supply
- 170 pmol m-2 d- 1 Hg, to the estuary, assuming a rain Hg, concentration of 50 pM and
10 cm of rain per month. Dry deposition would
contribute a comparable flux (130 pmol rnd2
d- l; 0.5 cm s- l depositional velocity, 60 pg
m-3 atmospheric particulate Hg, Fitzgerald et
al. 199 1). Atmospheric flux is the net source
of Hg to the system, assuming that stream fluxes into and out of the lower basin are similar.
The sinks for Hg are gas exchange (100-200
pmol m-2 d-l) and particulate deposition to
the sediment (Fig. 7). Net sedimentation is estimated, by difference, at 100-200 pmol me2
d-l. Partic u 1at e scavenging and sinking removes Hg from the bottom of the mixed layer
at a rate similar to the net sedimentation flux.
Thus, Hg” formation and evasion and particulate scavenging remove Hg, from the mixed
layer at similar rates and are the predominant
removal mechanisms for mixed-layer Hg,.
There is little formation of MMHg in this region and Hg” formation is the predominant
reaction consuming Hg, in the oxic region of
the estuary (Fig. 7). These processes also predominate in the epilimnion of the Wisconsin
lakes and in the mixed layer of equatorial Pacific Ocean (Mason 199 1).
Concentrations of MMHg, and MMHg, are
low in the mixed layer (upper 2 m), but some
methylation could occur in this region. The
principal external source of MMHg is atmospheric deposition, which would contribute
- l-2 pmol m-2 d- 1 MMHg, assuming concentrations in wet and dry deposition of the
same order as found for the Pettaquamscutt
estuary and in northern Wisconsin (Fitzgerald
et al. 199 1; Mason and Vandal unpubl. data).
Any MMHgD that diffuses into the mixed layer
from the pycnocline would be removed by particulate uptake or by demethylation. Demethylation ,could be important in the upper reaches of the pycnocline and in the mixed layer.
However, there is no strong MMHg,, gradient
between 2 and 4 m, indicating that the diffusive flux of MMHg, out of the pycnocline region is small. In addition, the magnitude and
position of the peak in MMHg, suggests that
most of the MMHgD formed in the pycnocline
region is scavenged from solution in the thermocline. Thus, the flux of MMHgr, out of the
pycnocline is relatively small and, as a result,
demethylation of MMHgD in the mixed layer
is not an important process.
Hg is transported downward by diffusion and
particulate sinking to the oxic-anoxic transition zone where Hg released as particulate is
remineralized. Methylation occurs in this region (Fig. 8) and is the principal reaction of
Hg, there. In the anoxic region there is little
Hg in the Pettaquamscutt
Atml>spheric
deposi
1239
ion
Mixed
layer
<
Pycnocline
Diffusion
Anoxic
zone
Diffueion
’ Hgtll)
<
V
Hg
c
.
P
Fig. 8. Model of MMHg biogeochemical cycling in the Pettaquamscutt estuary.
formation of MMHg and HgO. Particulate
scavenging and dissolution result in rapid cycling of both ionic Hg and MMHg in this zone.
Sinking particulate matter is remineralized in
the lower reaches of the pycnocline, but at
higher sulfide concentrations FeS precipitates,
scavenging Hg-S species from solution. Particulate dissolution maintains low MMHg, in
the upper reaches of the anoxic zone, and precipitation removes MMHg in the deeper waters. Sulfide complexation helps maintain the
relatively high MMHgD concentration and a
low Hg, concentration in this region. The generally low concentration of Hg, throughout the
water column is thus a result of continuous
removal by particulate scavenging and of conversion into MMHg and HgO.
There is little diffusion of MMHg, from the
sediments into anoxic waters, and MMHgP
concentrations generally decrease with depth
throughout the anoxic zone. Demethylation is
not a significant MMHg removal process below the mixed layer and particulate sinking is
the ultimate removal mechanism, even though
MMHg, concentrations are low. The flux to
the sediment is estimated at lo-20 pmol m-2
d-l (10% of the total flux).
The fluxes estimated for the Pettaquamscutt
estuary are similar to those calculated for Little
Rock Lake (Fitzgerald et al. 199 1). Atmospheric deposition does not provide sufficient
MMHg, the flux being equivalent to - 1% of
the net MMHg formed in the Pettaquamscutt
estuary. Both Hg” and MMHg are formed in
this basin and calculations show that demethylation is not an important sink for MMHg in
this system. Most of the MMHg formed is finally removed by deposition to the sediment.
1240
Mason et al.
References
BERMAN,M., T. CHASE,AND R. BARTHA. 1990. Carbon
flow in mercury biomethylation by Desulfovibrio desulfuricam. Appl. Environ. Microbial. 56: 298-300.
BLOOM, N. S. In press. Inlhrence of analytical conditions
on the observed “labile ionic mercury” concentrations in natural freshwaters. In C. J. Watras and J.
W. Huckabee [eds.], Mercury as a global pollutant.
Lewis.
-.
1989. Determinationofpicogram levels ofmethylmercury by aqueous phase ethylation, followed by
cryogenic gas chromatography with atomic fluorescence detection. Can. J. Fish. Aquat. Sci. 46: 113l1140.
-,
AND W. F. FITZGERALD. 1988. Determination of
volatile mercury species at the picogram level by low
temperature gas chromatography with cold-vapour
atomic fluorescence detector. Anal. Chim. Acta 208:
151-161.
-,
AND C. J. WATRAS. 1989. Seasonal and vertical
variability in the mercury speciation of an acidified
Wisconsin lake, p. 349-352. In J.-P. Vernet [ed.],
Heavy metals in the environment. V. 2. CEP Consultants.
-AND J. P. HURLEY. 199 1. Impact of acidifiiation oh the methylmercury cycle of remote seepage lakes. Water Air Soil Pollut. 56: 477-491.
COMPEAU, G., AND R. BARTHA. 1985. Sulfate-reducing
bacteria: Principal methylators of mercury in anoxic
estuarine sediment. Appl. Environ. Microbial. 50: 498502.
COSSA, D., C. GOBEIL, AND P. COURAU. 1988. Dissolved
mercury bchaviour in the Saint Lawrence estuary.
Estuarine Coastal Shelf Sci. 26: 227-230.
AND J.-M. MARTIN. 199 1. Mercury in the Rhone
delta and adjacent marine areas. Mar. Chem. 36: 29 l302.
-,
AND J. NoBL. 1987. Concentrations of mercury
in near shore surface waters of the Bay of Biscay and
in the Gironde estuary. Mar. Chem. 20: 389-396.
DONAGHAY, P. L., H. M. RINES, AND J. McN. SIEBURTH.
1992. Simultaneous sampling of fine scale biological,
chemical and physical structure in stratified waters.
Arch. Hydrobiol. 36: 97-108.
DYRSSEN, D. 1989. Sulphide complexation in surface
seawater. Mar. Chem. 24: 143-153.
-,
AND K. KREMLING. 1990. Increasing hydrogen
sulphide concentration and trace metal behavior in
the anoxic Baltic waters. Mar. Chem. 30: 193-204.
EREL, Y., AND J. J. MORGAN. 199 1. The effect of surface
reactions on the relative abundances of trace metals
in deep ocean waters. Geochim. Cosmochim. Acta
55: 1807-1813.
FAGERSTROM, T., AND A. JERNEL~V. 197 1. Formation of
mcthylmercury from pure mercuric sulphide in aerobic organic sediment. Water Res. 5: 121-122.
FITZGERALD, W. F., AND T. W. CLARKSON. 1991. Mercury and monomethylmercury: Present and future
concerns. Environ. Health Perspect. 96: 159-l 66.
-,
G. V. VANDAL, AND R. P. MASON. 1991. Atmospheric cycling and air-water exchange of mercury
over mid-continental lacustrine regions. Water Air
Soil Pollut. 56: 745-767.
3 AND C. J. WATRAS. 1989. Mercury in surficial
waters of rural Wisconsin lakes. Sci. Total Environ.
87188: 223-232.
GAINES, A. G., JR. 1975. Papers on the geomorphology,
hydrography and geochemistry of the Pettaquamscutt
estuary. Ph.D. thesis. Univ. Rhode Island. 260 p.
GILL, G. A., AND W. F. FITZGERALD. 1987. Picomolar
mercury measurement in seawater and other materials using stannous chloride reduction and two-stage
gold amalgamation with gas phase detection. Mar.
Chem. 20: 227-243.
GILMOUR, C. G., AND E. A. HENRY. 199 1. Mercury methylation in aquatic systems affected by acid deposition.
Environ. Pollut. 71: 13l-l 70.
HARALDSSON, C., AND S. WESTERLUND. 1988. Trace metals in the water column of the Black Sea and Framvaren Fjord. Mar. Chem. 23: 417-424.
HONEYMAN, B. D., AND P. H. SANTSCHI. 1988. Metals
in aquatic systems. Environ. Sci. Technol. 22: 862872.
HURLEY, J. P., AND C. J. WATRAS. 199 1. Identification
of bacteriochlorophylls in lakes via reverse-phase
HPLC. Limnol. Oceanogr. 56: 307-315.
-,
AND N. S. BLOOM. 199 1. Mercury cycling
in ‘northern Wisconsin seepagelakes: The role of particulate matter in vertical transport. Water Air Soil
Pollut. 36: 307-3 15.
IVERFELDT, A. 1988. Mercury in the Norwegian fiord
Framvaren. Mar. Chem. 23: 441-456.
KIM, J. P., AND W. F. FITZGERALD. 1986. Sea-air partitioning of mercury in the equatorial Pacific Ocean.
Science 231: 113 l-l 133.
LEE, Y.-H., AND H. HULTBERG. 1990. Methylmercury in
some Swedish surface waters. Environ. Toxicol. Chem.
9: 833-841.
LINDQVIST, O., AND OTHERS. 199 1. Mercury in the Swedish environment-recent research on causes, consequences and corrective methods. Water Air Soil Polht. 55: 143-177.
Lu, X., W. K. JOHNSON, AND C. S. WONG. 1986. Seasonal
replenishment of mercury in a coastal fjord by its
intermittent anoxicity. Mar. Pollut. Bull. 17: 263-267.
MADIGAN, M. T. 1988. Microbiology, physiology, and
ecology of phototrophic bacteria, p. 39-l 11. In A. J.
B. Zehnder [ed.], Biology of anaerobic microorganisms. Wiley.
MANTOURA, R. F. C., A. DICKSON, AND J. P. RILEY. 1978.
The complexation of metals with humic materials in
natural waters. Estuarine Coastal Shelf Sci. 6: 387. 408.
MASON, R. P. 199 1. The chemistry of mercury in the
equatorial Pacific Ocean. Ph.D. thesis, Univ. Connecticut. 305 p.
AND W. F. FITZGERALD. 1990. Alkylmercury species in the equatorial Pacific. Nature 347: 457-459.
1991. Mercury speciation in open
-,ANDocean waters. Water Air Soil Pollut. 56: 779-789.
MILLS, G. L., G. S. DOUGLAS, AND J. G. QUINN. 1989.
Dissolved organic copper isolated by C,, reverse-phase
extraction in an anoxic basin located in the Pettaquamscutt estuary. Mar. Chem. 26: 277-288.
SARMIENTO, J. L., H. W. FEELY, W. S. MOORE, A. E. BAINBRIDGE, AND W. S. BROECKER. 1976. The relationship between vertical eddy diffusion and buoyancy
-
Hg in the Pettaquamscutt
gradient in the deep sea. Earth Planet. Sci. Lett. 32:
357-370.
SCRANTON, M. I., P. GRILL, M. A. DEANGELIS, P. L.
DONAGHAY, AND J. McN. SIEBURTH. In press. The
importance of episodic events in controlling flux of
methane from the anoxic zone of the lower basin of
the Pettaquamscutt River estuary. Global Biogeothem. Cycles.
TOPPING, G., AND I. M. DAVIES. 198 1. Methylmercury
production in the marine water column. Nature 290:
243-244.
1241
VANDAL, G. M., W. F. FITZGERALD, AND R. P. MASON.
199 1. Cycling of volatile mercury in temperate lakes.
Water Air Soil Pollut. 56: 791403.
WINFREY, M. R., AND J. W. M. RUDD. 1990. Environmental factors affecting the formation of methylmercury in low pH lakes. Environ. Toxicol. Chem. 9: 853869.
Submitted: 20 March 1992
Accepted: 15 December 1992
Revised: 15 January 1993