DSpace at VNU: Influence of salinity intrusion on the speciation and partitioning of mercury in the Mekong River Delta

Available online at www.sciencedirect.com

Geochimica et Cosmochimica Acta 106 (2013) 379–390
www.elsevier.com/locate/gca

Influence of salinity intrusion on the speciation and partitioning
of mercury in the Mekong River Delta
Seam Noh a, Mijin Choi a, Eunhee Kim a, Nguyen Phuoc Dan b, Bui Xuan Thanh b,
Nguyen Thi Van Ha c, Suthipong Sthiannopkao d, Seunghee Han a,⇑
a

School of Environmental Science and Engineering, Gwangju Institute of Science and Technology (GIST), Gwangju 500-712, Republic of Korea
b
Faculty of Environment, Ho Chi Minh City University of Technology, District 10, Ho Chi Minh City, Viet Nam
c
Faculty of Environment, Ho Chi Minh City University of Natural Resources and Environment, Tan Binh District, Ho Chi Minh City, Viet Nam
d
Department of Environmental Engineering, Dong-A University, Busan 604-714, Republic of Korea
Received 26 February 2012; accepted in revised form 11 December 2012; available online 29 December 2012

Abstract
The lower Mekong and Saigon River Basins are dominated by distinctive monsoon seasons, dry and rainy seasons. Most
of the Mekong River is a freshwater region during the rainy season, whereas during the dry season, salt water intrudes
approximately 70 km inland. To understand the role of salinity intrusion controlling Hg behavior in the Mekong and Saigon
River Basins, Hg and monomethylmercury (MMHg) in surface water and sediment of the Mekong River and in sediment of
the Saigon River were investigated in the dry season. Sediment Hg distribution, ranging from 0.12 to 0.76 nmol gÀ1, was
mainly controlled by organic carbon distribution in the Mekong River; however, the location of point sources was more
important in the Saigon River (0.21–0.65 nmol gÀ1). The MMHg concentrations in Mekong (0.16–6.1 pmol gÀ1) and Saigon
(0.70–8.7 pmol gÀ1) sediment typically showed significant increases in the estuarine head, with sharp increases of acid volatile
sulfide. Unfiltered Hg (4.6–222 pM) and filtered Hg (1.2–14 pM) in the Mekong River increased in the estuarine zone due to
enhanced particle loads. Conversely, unfiltered MMHg (0.056–0.39 pM) and filtered MMHg (0.020–0.17 pM) was similar

between freshwater and estuarine zones, which was associated with mixing dilution of particulate MMHg by organic- and
MMHg-depleted resuspended sediment. Partitioning of Hg between water and suspended particle showed tight correlation
with the partitioning of organic carbon across study sites, while that of MMHg implied influences of chloride: enhanced chloride in addition to organic matter depletion decreased particulate MMHg in the estuarine zone. Primary production was an
important determinant of inter-annual variation of particulate Hg and sediment MMHg. The bloom year showed relatively
low particulate Hg with low C/N ratio, indicating biodilution of Hg. In contrast, the percentage of MMHg in sediment
increased significantly in the bloom year, likely due to greater availability of metabolizable fresh organic matter. The overall
results emphasize that Hg behavior in the lower Mekong River Basin is strongly connected to the local monsoon climate, via
alterations in particle loads, biological productivity, and availability of sulfate, chloride and organic matter.
Ó 2012 Elsevier Ltd. All rights reserved.

1. INTRODUCTION
Interactions between Hg and organic matter influence
the transport and bioavailability of Hg in riverine and

⇑ Corresponding author. Tel.: +82 62 7152438; fax: +82 62
7152434.
E-mail address: (S. Han).

0016-7037/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved.
/>
estuarine environments (Laurier et al., 2003; Choe et al.,
2003; Conaway et al., 2003; Han et al., 2007; Lee et al.,
2011). Dissolved Hg distribution across watersheds is explained by dissolved organic carbon (DOC) distribution
in a number of river systems (Peckenham et al., 2003;
Schuster et al., 2008; Brigham et al., 2009). In estuarine
systems, complexation of Hg with dissolved organic matter, coupled with colloidal coagulation, is reported to
influence estuarine mixing behavior (Stordal et al., 1996;


380

S. Noh et al. / Geochimica et Cosmochimica Acta 106 (2013) 379–390

Choe et al., 2003; Lee et al., 2011) and bioavailability of
Hg (Pan and Wang, 2004; Zhong and Wang, 2009).
Monomethylmercury (MMHg) is a toxic form of Hg that
biomagnifies in aquatic food chains (Chen et al., 2008;
Gantner et al., 2009). Previous studies note that biogeochemical factors (e.g., sulfate, sulfide, and organic matter
contents) are critical in the production of MMHg in estuarine sediments (Hammerschmidt et al., 2004; Hammerschmidt and Fitzgerald, 2006; Han et al., 2007). For
example, increasing Hg(II) methylation rate along with
increasing salinity is shown in estuarine sediments to be
associated with sulfate availability to sulfate-reducing bacteria (Hollweg et al., 2009).
The Mekong River originates at an elevation of about
5100 m in the Tibetan Plateau of western China and flows
southward through China, Myanmar, Laos, Thailand,
Cambodia, and Vietnam, before discharging into the
South China Sea (Edwin, 2009). It has a total length of
4800 km and drains an area of 795,000 km2, with a mean
annual water discharge of 470 km3 yrÀ1, making it the longest river in Southeast Asia (Dai and Trenberth, 2002).
The river flows over rock for about 80% of its length before it enters the alluvial plain of Cambodia and Vietnam
(Xue et al., 2010, 2011). The Mekong River Delta in
southern Vietnam is composed of Holocene alluvial sediments that were rapidly deposited beginning 8000 yr BP
(144 Â 106 ton yrÀ1; Xue et al., 2010, 2011). The delta is
an essential component of life for millions of South Asians
residing in the vicinity of the river and its tributaries, providing the main freshwater source for urban, industrial,
agricultural, and fishery uses (Osborne, 2000; Baran
et al., 2007). Despite this importance, surprisingly little
information is available regarding Hg levels in the Mekong River as compared to other major rivers (e.g., the
Amazon and Nile Rivers).
The Saigon River originates from Phum Daung in

southeastern Cambodia, flows south for approximately
225 km, and empties into the Nha Be River, which discharges into the South China Sea, 20 km north of the Mekong Delta (Lambert et al., 2010). The Saigon River is
joined by the Ben Cat River just upstream of Ho Chi Minh
City, and it encircles Ho Chi Minh City, the largest city in
Vietnam. Since the 1980s, Ho Chi Minh City has experienced continually increasing population density and rapid
industrial growth, which have negatively affected the environmental quality of the urban river (Thuy et al., 2007; Vo,
2007). Large volumes of untreated domestic and industrial
wastewater, along with substances from accidental spills,
are released directly into the river (Minh et al., 2007).
Monitoring of metals pollution in the Saigon River indicates that surface river-water and sediments are severely
contaminated with Cd, Cr, Cu and Zn (Thuy et al.,
2007); nevertheless, no literature is available regarding levels of Hg pollution in the water and sediment of the Saigon
River.
The lower Mekong River and Saigon River Basins are
dominated by distinctive monsoon seasons. The dry season
occurs from November to April, with an average discharge
rate of 2000 m3 sÀ1, and the rainy season occurs from May
to October, with an average discharge rate of 40,000 m3 sÀ1

for the Mekong River (Hoa et al., 2007). While most of the
Mekong Delta is a freshwater region during the rainy season, salt water intrudes approximately 70 km inland in
the dry season, with vertical stratification in salinity
(Wolanski et al., 1996, 1998; Cenci and Martin, 2004). In
the wet season, salinity intrusion is observed near the river
mouth, but this does not extend more than a few kilometers
inland (Wolanski et al., 1996). In contrast to the Mekong
River, the Saigon River has a constant flow rate
(54 m3 sÀ1; People’s Committee of Ho Chi Minh City,
2006) throughout the year, due to flow regulation from
the Dau Tieng Reservoir (George et al., 2004). Salinity

intrusion is observed from the lower Saigon River, approximately 10 km inland from the Nha Be River
(www.eng.hochiminhcity.gov.vn).
Salinity intrusion influences the speciation and bioavailability of Hg in low-discharge rivers (Chiffoleau et al.,
1994; Laurier et al., 2003; Covelli et al., 2006). Like permanent estuaries, interactions between Hg and organic
matter have primary importance in the transport and bioavailability of Hg in these rivers (Laurier et al., 2003;
Turner et al., 2004). Hg–chloro complexation could increase the mobility of Hg, even though the majority of
Hg may remain in organic fraction (Ramalhosa et al.,
2005). Salinity intrusion is often accompanied by formation of a high turbidity zone as a result of intense hydrodynamic energy of tidal currents (Laurier et al., 2003;
Covelli et al., 2006). In a high turbidity zone, the quantity
and quality of particulate organic matter influence redistribution of Hg between water and suspended particles,
which may cause changes in Hg solubility and bioavailability (Laurier et al., 2003; Turner et al., 2004; Covelli
et al., 2006; Cana´rio et al., 2008). A high turbidity zone
often provides an ideal site for diagenetic processes of
trace metals, in association with increased microbial activity and flocculation of colloidal particles (Roth et al.,
2001; Laurier et al., 2003; Covelli et al., 2006). Increased
MMHg concentration in the high turbidity zone at the
mouth of the Isonzo River is considered to result from intense microbial methylation in the low-oxygen bottomwater (Covelli et al., 2006).
In the current study, we aim to understand the role of
salinity intrusion on Hg speciation in riverine water and
sediment. We hypothesize that: (1) salinity intrusion increases sediment MMHg production due to the enhanced
availability of sulfate to Hg(II)-methylating microbes; (2)
salinity intrusion increases particle solubility of Hg and
MMHg due to chloride complexation and/or particulate
organic matter dilution by resuspended sediment; and (3)
salinity intrusion increases water column MMHg concentration, as a result of intense microbial activity in the high
turbidity zone. To test these hypotheses, we examine Hg
and MMHg distributions in surface river water, suspended
particles, and sediments in relation to relevant biogeochemical variables (e.g., salinity, suspended particulate
matter (SPM), sulfate, chlorophyll-a (Chl-a), DOC, particulate organic carbon (POC), and particulate nitrogen
(PN)); and of sediment (e.g., total organic carbon

(TOC), total nitrogen (TN) and acid volatile sulfide
(AVS)).


S. Noh et al. / Geochimica et Cosmochimica Acta 106 (2013) 379–390

381

2. MATERIALS AND METHODS
2.1. Sample collection and pre-treatment
Water and sediment samples were collected from 14 sites
in March 2010 and seven sites in April 2011 along the Tien
(Fig. 1), and from eight sites along the Saigon River in
March 2010 (Fig. 2). Ultra-clean sample handling protocols
were used to prevent sample contamination. Unfiltered surface water samples were collected from a depth of approximately 0.5 m in acid-cleaned Teflon bottles, using a
peristaltic pump system equipped with Teflon tubing. Filtered surface water samples were collected using the same
method with polyethersulfone filter capsules (MilliporeÒ,
0.45 lm) connected to the tubing outlet. At each site, three
independent samples were collected in Teflon, polyethylene,
and borosilicate glass bottles. The Teflon bottles of unfiltered and filtered water for determination of Hg and
MMHg were preserved with high-purity HCl (0.2% v/v),
and stored at 4 °C within 12 h. Polyethylene bottles of unfiltered water were kept at 4 °C for measurements of SPM,
POC, PN, alumina, and chlorophyll-a concentrations.
Borosilicate glass vials of filtered water were collected for
determination of DOC and sulfate concentration. Filtered
water samples were collected only from the Tien River.
The ancillary parameters, e.g., pH, temperature, salinity,
and oxidation–reduction potential (ORP), were recorded
in situ using a multi-parameter probe (Thermo Scientific).
The unfiltered water samples for SPM, POC, PN, alumina, and chlorophyll-a were filtered in the laboratory

within 24 h of sampling. Glass fiber filters (WhatmanÒ,
GF/F) were used for particle collection for chlorophyll-a,
alumina, POC, and PN measurements, and 0.4 lm polycarbonate membranes were used for SPM measurements. The

Fig. 2. Map of sampling sites in Saigon River, Vietnam.

filters containing particles were dried immediately for measurement of SPM and alumina, and filters for chlorophylla, POC, and PN were frozen at À20 °C until analysis. Sediment samples were collected from the top-20 cm layer

Fig. 1. Map of sampling sites in Tien River, Vietnam.


382

S. Noh et al. / Geochimica et Cosmochimica Acta 106 (2013) 379–390

using a stainless steel grab sampler to determine Hg,
MMHg, and TOC; and TN and AVS concentrations. The
collected sediments were stored in sealed polyethylene bags
and frozen at À20 °C until analysis.

difference of duplicate analyses was typically 15 ± 14%
(n = 73), whereas certified reference material (ERMCC580, IRMM, 75 ± 4 lg kgÀ1) recovery averaged
96 ± 8% (n = 8).

2.2. Hg and MMHg analysis

2.3. Particle and sediment compositions

The Hg in water samples was analyzed following EPA
method 1631 (2002). Aqueous samples were oxidized with

bromine monochloride (BrCl) for a minimum of 12 h. After
oxidation, excess oxidant was destroyed with hydroxylamine hydrochloride solution prior to analysis. Divalent
Hg in these samples was reduced to elemental Hg by SnCl2
solution and the elemental Hg was trapped on gold traps.
The Hg0 released from gold traps by thermal desorption
was fed into a cold vapor atomic fluorescence spectrometer
(CVAFS; Model III, Brooks Rand). An acceptable calibration curve was achieved daily with an r2 of at least 0.99. The
relative differences of duplicate analyses averaged 11 ± 8%
(n = 144) and recovery of the matrix spike averaged
98 ± 25% (n = 7). Recovery of certified reference material
(European Commission, BCRÒ-579, coastal seawater,
1.9 ± 0.5 ng kgÀ1) averaged 105 ± 6% (n = 10).
Details of the analytical protocols for MMHg analysis
in an aqueous phase were as given by EPA method 1630
(2001). Acidified water samples were distilled with ammonium pyrrolidine dithiocarbamate (APDC) solution and
the distillates were then converted to gaseous MMHg by
aqueous-phase ethylation using tetraethylborate solution.
The volatile MMHg was then purged and trapped onto TenaxÒ traps, which were flash-heated in a nitrogen stream.
The released Hg species were thermally separated on a
GC column, then detected by CVAFS (Model III, Brooks
Rand). An acceptable calibration curve was achieved daily
with an r2 of at least 0.99. The relative difference of duplicate analyses was typically 18 ± 5% (n = 52), whereas the
matrix spike and certified reference material (ERMCE464, IRMM, 5.5 ± 0.2 mg kgÀ1) recovery averaged
99 ± 6% (n = 12) and 99 ± 6% (n = 12), respectively.
The concentrations of particulate Hg (PHg) and particulate MMHg (PMMHg) were determined from the differences between the unfiltered Hg (or MMHg) and filtered
Hg (or MMHg), normalized to the SPM concentration.
The Hg found in sediment was analyzed following the
appendix to EPA method 1631 (2001). The Hg in sediment
(0.5–1.5 g) was digested overnight with 10 ml of aqua regia.
Sediment digests were then diluted with 0.07 N BrCl solution and used for Hg determination, using the same method

as for aqueous Hg samples. The relative difference of
duplicate analyses was typically 16 ± 24% (n = 48), whereas
certified reference material (ERM-CC580, IRMM, 132 ±
3 mg kgÀ1) recovery ranges were 92 ± 6% (n = 12).
The MMHg in sediment was analyzed following the procedure described by Choe and Gill (2003). The sediment
MMHg (0.5–1.0 g) was digested with acidic potassium bromide solution and extracted into methylene chloride. An
aliquot of methylene chloride was then back-extracted to
Milli-Q water by purging out methylene chloride with nitrogen gas. The extracted MMHg was measured using the
same method described for aqueous samples. The relative

The SPM was measured by filtering 0.2 L of water
through a pre-weighed polycarbonate membrane (WhatmanÒ, 0.4 lm) shortly after sample collection. Particles
were dried at 60 °C for 12 h, and the SPM loads were then
weighed. For determination of organic carbon and nitrogen
in SPM and sediment, the GF/F filters bearing the SPM
and approximately 10 mg of sediment were freeze-dried.
Freeze-dried samples were acidified with HCl solution to remove inorganic carbon, then measured with an elemental
analyzer interfaced to a continuous flow isotope ratio mass
spectrometer (EA-IRMS, ThermoQuest). Alumina on filters was analyzed with a wavelength dispersive X-ray fluorescence spectrometer (Axios Mineral, PANalytical). The
samples for chlorophyll-a were obtained by filtering 0.1–
0.4 L of water through GF/F filters. Phytoplankton pigments retained on the GF/F filters were extracted in 90%
acetone (Liu et al., 2007). The chlorophyll-a in the extracts
was measured at 750, 665, 645, 630 and 480 nm wavelengths with an ultraviolet–visible spectrophotometer (Optizen, Mecasys). DOC and sulfate concentrations in filtered
water samples were measured by TOC analyzer (Multi N/C
3100, Analytik Jena) and ion chromatography (DX-120,
Dionex), respectively. For determination of AVS, approximately 10 g of sediment sample was acidified with 20 ml of
6.0 M deoxygenated HCl and held under a nitrogen gas
flow for 2 h. Volatilized sulfide was collected in traps filled
with sulfide antioxidant buffer (SAOB; EPA, 1996), then
measured with a sulfide electrode (Kim et al., 2006).

Sigma PlotÒ, version 11.2 (Systat software, Inc.) was
used for all statistical analyses. Linear regression analysis
yielded the coefficient of determination (r2) and significant
probability (p). Differences between two independent
groups were determined using one-way ANOVA. All statistical results were reported as significant at a level of
p < 0.05. Linear correlation analyses yielded the Pearson’s
correlation coefficient (r) for parameters passing the normality test (Shapiro–Wilks normality test). The results were
considered statistically significant if p-values were less than
0.05.
3. RESULTS
3.1. Geochemical settings
In the Tien River, the mean water temperature across
sampling sites was 31 ± 0.9 °C in 2010 and 30 ± 0.8 °C in
2011, and mean pH was 7.8 ± 0.1 in 2010 and 7.4 ± 0.3
in 2011. Saline intrusion was observed from 0 to 50 km
from the coast in 2010 and from 0 to 20 km in 2011 (Supporting Information, Table S1). In the brackish zone of
the Tien River, the residence time of saline bottom-water
can be quite long, as indicated by lower ORP. High concentrations of SPM (>80 mg LÀ1) were found from the brack-


76 ± 65
23–160
43 ± 41
14–90
21 ± 1.4
19–22
2862 ± 3064
314–6263
5.2 ± 0.48
4.7–5.8

6.2 ± 2.4
4.5–9.0
0.92 ± 0.31
0.48–1.2
0.64 ± 0.40
0.22–1.0
4.0 ± 1.1
2.4–4.8
2.9 ± 1.1
1.7–4.0
ND

3.4 ± 0.50
3.0–4.1
4.0 ± 0.25
3.8–4.3
ND

ND

24 ± 12
16–43
185 ± 106
81–293
ND

ND: not determined.

Mean ± SD
Range

Mean ± SD
Range
19–21 Brackish

10–14 Brackish

7–9 Lower freshwater

15–18 Lower freshwater

9.4 ± 3.8
4.1–14
7.1 ± 3.3
4.0–11
4.0 ± 1.1
2.8–5.3
14 ± 0.55
13–15
15 ± 1.7
13–16
946 ± 693
91–1918
5.8 ± 0.31
5.4–6.2
6.2 ± 0.24
5.9–6.4
9.6 ± 1.9
7.6–12.6
1.2 ± 0.45
0.61–1.8

0.72 ± 0.21
0.50–0.92
0.22 ± 0.050
0.16–0.29
6.0 ± 2.2
3.3–9.3
3.8 ± 1.2
2.7–5.0
1.7 ± 0.17
1.6–1.9
2.8 ± 1.9
1.4–6.4
2.0 ± 0.51
1.5–2.5
2.8 ± 0.76
2.1–4.1
1.8 ± 0.19
1.6–2.0
1.7 ± 0.21
1.5–1.9
2.4 ± 0.084
2.3–2.5
13 ± 5.0
7.8–20
16 ± 3.9
12–18
289 ± 195
82–517
1–6 Upper freshwater

Mean ± SD
Range
Mean ± SD
Range
Mean ± SD
Range

152 ± 27
123–195
162 ± 56
112–222
104 ± 35
61–154

Sulfate
(mg LÀ1)
C/N
(mol/mol)
PN
(%)
POC
(%)
DOC
(mg LÀ1)
Alumina
(mmol gÀ1)
SPM
(mg LÀ1)

383

Tien 2011

The Hg concentrations found in Tien River sediment
ranged from 0.32 to 0.76 nmol gÀ1 dry weight (mean
0.50 ± 0.14 nmol gÀ1) in 2010 and from 0.12 to
0.24 nmol gÀ1 dry weight (mean 0.18 ± 0.040 nmol gÀ1) in
2011 (Table 2); in the Saigon River, it ranged from 0.21

Tien 2010

3.3. Total Hg and MMHg in sediment

ORP
(mV)

In the Tien River, TOC in sediment ranged from 2.2% to
3.9% (mean 2.7 ± 0.58%) in 2010, and from 1.2% to 2.4%
(mean 1.7 ± 0.5%) in 2011. In the Saigon River, TOC in
sediment ranged from 2.6% to 3.7% (mean 3.2 ± 0.32%)
(Table 2). Little lateral variation of TOC, TN, and C/N ratios were noted in the Tien and Saigon Rivers (Supporting
Information, Table S3). Temporally, atomic C/N ratios in
the Tien River were significantly (p < 0.05) lower in 2011
(range 3.4–6.5) than in 2010 (range 9.6–15).
AVS in sediment of the Tien River ranged from 0.053 to
6.5 lmol gÀ1 in 2010 and from 0.20 to 20 lmol gÀ1 in 2011
(Table 2). Mean AVS concentration was significantly higher (p < 0.05) in the brackish zone compared to the freshwater zone in 2010. The same distribution pattern was found
in the Saigon River, where AVS ranged from 0.14 to
0.45 lmol gÀ1 in the freshwater zone, and from 5.7 to
17 lmol gÀ1 in the brackish zone (p < 0.05). The primary

factors that control AVS in river and estuarine sediments
are the availabilities of reactive Fe(II), dissolved sulfate,
and metabolizable organic carbon (Morse et al., 2007). As
the concentrations of TOC and TN, and C/N ratio were
consistent across sites, increased AVS at the saline zone
could be attributable to enhanced sulfate availability.

Site

3.2. Sediment characteristics

Table 1
Data for ORP, SPM, Alumina, DOC, POC, PN, atomic C/N, sulfate and chlorophyll-a (Chl-a) in surface waters of the Tien River, collected in March 2010 and 2011.

ish zone, and were associated with hydrodynamic energy of
tidal currents: the Mekong River Delta has 3.5-m semidiurnal tides from the South China Sea and irregular 1-m diurnal tides from the Gulf of Thailand (Hoa et al., 2007). In
the Saigon River, the mean water temperature and pH for
2010 were 30 ± 0.9 °C and 7.7 ± 0.3, respectively, and saline intrusion was observed up to 10 km from the Nha Be
River (Supporting Information, Table S2). Like the Tien
River, the brackish zone showed lower ORP than the freshwater zone, and a high turbidity zone was found at sites 7
and 8.
Concentration and compositional characteristics of the
SPM (i.e., alumina, POC, PN, and atomic C/N ratio) were
compared between freshwater and brackish zones of the
Tien River (Table 1). Alumina content, used as a proxy
for fine particles, significantly (p < 0.05) increased in the
saline zone, supporting the periodic occurrence of sediment
resuspension. Levels of POC varied from 1.6% to 9.3%
(mean 4.0 ± 2.4%) in 2010 and from 1.7% to 4.8% (mean
3.5 ± 1.2%) in 2011. POC decreased in the brackish zone

in both years, implying dilution of fluvial particles by organic-depleted resuspended sediment. The PN showed a
similar trend to the POC. The strong correlation between
POC and PN (r2 = 0.95, p < 0.05, linear regression) suggests that N is predominantly bound to particulate organic
matter.

Chl-a
(lg LÀ1)

S. Noh et al. / Geochimica et Cosmochimica Acta 106 (2013) 379–390


0.40 ± 0.063
0.33–0.46
1.2 ± 0.78
0.75–2.6
1.5 ± 1.0
0.70–2.6
5.0 ± 2.1
3.2–8.7
0.38 ± 0.24
0.21–0.65
0.47 ± 0.11
0.34–0.61
13 ± 3.2
9.1–15
11 ± 0.84
10–12
6–8 Brackish

Mean ± SD

Range
Mean ± SD
Range
1–5 Freshwater
Saigon 2010

19–21 Brackish

3.1 ± 0.41
2.6–3.4
3.3 ± 0.29
2.9–3.7

0.29 ± 0.041
0.25–0.33
0.35 ± 0.017
0.33–0.38

0.26 ± 0.17
0.14–0.45
13 ± 4.6
5.7–17

1.1 ± 1.1
0.24–2.7
0.85 ± 0.51
0.40–1.4
2.2 ± 2.1
0.57–5.2
1.2 ± 0.64

0.81–2.0
0.20 ± 0.030
0.17–0.24
0.15 ± 0.040
0.12–0.20
5.2 ± 1.3
3.7–6.5
3.9 ± 0.44
3.3–4.1
Mean ± SD
Range
Mean ± SD
Range
15–18 Lower freshwater
Tien 2011

10–14 Brackish

7–9 Lower freshwater

1.9 ± 0.61
1.2–2.4
1.4 ± 0.17
1.3–1.6

0.43 ± 0.046
0.36–0.47
0.44 ± 0.041
0.39–0.47

6.3 ± 9.3
0.20–20
2.2 ± 2.4
0.81–4.9

0.64 ± 0.28
0.16–0.99
1.4 ± 1.5
0.33–3.1
1.7 ± 2.5
0.44–6.1
0.43 ± 0.11
0.32–0.63
0.54 ± 0.14
0.46–0.70
0.55 ± 0.15
0.35–0.76
12 ± 2.1
10–15
10 ± 0.7
10–11
10 ± 0.9
10–12
Tien 2010

1–6 Upper freshwater

Mean ± SD
Range
Mean ± SD

Range
Mean ± SD
Range

2.7 ± 0.14
2.5–2.9
3.1 ± 1.2
2.3–3.9
2.6 ± 0.70
2.2–3.8

0.26 ± 0.041
0.22–0.31
0.34 ± 0.11
0.27–0.42
0.31 ± 0.091
0.24–0.47

0.14 ± 0.11
0.053–0.31
1.6 ± 1.7
0.47–3.6
3.6 ± 2.6
1.4–6.5

MMHg
(pmol gÀ1)
Hg
(nmol gÀ1)
AVS

(lmol gÀ1)
C/N
(mol/mol)
TN
(%)
TOC
(%)
Site

Table 2
Data for TOC, TN, atomic C/N, AVS, Hg, MMHg and %MMHg in sediment of the Tien and Saigon Rivers, collected in March 2010 and 2011.

0.16 ± 0.089
0.051–0.31
0.22 ± 0.20
0.072–0.44
0.25 ± 0.30
0.085–0.79

S. Noh et al. / Geochimica et Cosmochimica Acta 106 (2013) 379–390

%MMHg

384

to 0.65 nmol gÀ1 (mean 0.44 ± 0.16 nmol gÀ1). These
ranges are comparable to those reported for urban rivers
and estuaries moderately contaminated with Hg: the Patuxent River (0.29–0.79 nmol gÀ1; Benoit et al., 1998) and Bay
of Fundy (0.050–0.70 nmol gÀ1; Sunderland et al., 2006).
There was no statistical (p > 0.05) difference between the

freshwater and saline zones in the Tien River, with weak increases at sites 9 and 10, located close to My Tho City (Supporting Information, Table S3). The Saigon River also
showed relatively constant sediment Hg, with weak peaks
at sites 1 and 7, located near Dau Tieng Reservoir and
Ho Chi Minh City, respectively.
In the Tien River, sediment MMHg ranged from 0.16 to
6.1 pmol gÀ1 dry weight (mean 1.2 ± 1.6 pmol gÀ1) in 2010,
and from 0.57 to 5.2 pmol gÀ1 dry weight (mean
1.8 ± 1.6 pmol gÀ1) in 2011 (Table 2). In the Saigon River,
sediment MMHg ranged from 0.70 to 8.7 pmol gÀ1 (mean
3.7 ± 2.5 pmol gÀ1) in 2010, comparable to data reported
for the Patuxent River (0.49–4.0 pmol gÀ1; Benoit et al.,
1998), the Bay of Fundy (0.25–7.38 pmol gÀ1; Sunderland
et al., 2006), and San Francisco Bay (0.5–5.0 pmol gÀ1;
Conaway et al., 2003). In the Tien River, strong MMHg
peaks were found near the estuarine head (sites 9, 10, and
18; Supporting Information, Table S3). In the Saigon River, MMHg concentrations were significantly (p < 0.05)
higher in the brackish zone than in the freshwater zone,
with a strong peak near the estuarine head, like the Tien
(site 4).
3.4. Total Hg and MMHg in surface water
In the Tien River, Hg levels in unfiltered river water
(UHg) ranged from 11 to 222 pM (mean 61 ± 65 pM) in
2010 and from 4.6 to 55 pM (mean 22 ± 23 pM) in 2011
(Fig. 3; Supporting Information, Table S4). These concentration ranges are similar or lower than those reported for
urban rivers in China and urban estuaries in North America moderately contaminated with Hg: the Yalujiang River
(154–344 pM; Zhang and Wong, 2007), East River (55–
244 pM; Liu et al., 2012), New York/New Jersey Harbor
(30–550 pM; Balcom et al., 2008), and San Francisco Bay
(0.73–440 pM; Conaway et al., 2003). Increased UHg levels
were found from the saline zone in both years. In the freshwater zone, a maximum peak of UHg was detected at site 6,

which might be a local runoff effect from Cao Lanh City.
Temporally, UHg was about four times higher in 2010 than
2011 in the lower-river and estuarine zone. DHg averaged
20% of UHg (4.7–44%), and spatial and temporal variations of DHg were similar to those of UHg. PHg levels were
decreased in the lower-river and estuarine zone compared
to the upper river (>100 km from coast), a major contrast
to the UHg and DHg distributions. Temporally, PHg was
higher in 2010 than in 2011.
In the Tien River, MMHg concentrations in unfiltered
water (UMMHg) ranged from 0.056 to 0.39 pM (mean
0.20 ± 0.085 pM) in 2010 and 0.14 to 0.28 pM (mean
0.20 ± 0.047 pM) in 2011 (Fig. 3; Supporting Information,
Table S4). These ranges are within the range previously reported for Hg-contaminated urban rivers and estuaries:
Patuxent River (0.05–1.2 pM; Benoit et al., 1998) and San


S. Noh et al. / Geochimica et Cosmochimica Acta 106 (2013) 379–390

385

Tien River ranged from 0.020 to 0.17 pM (mean
0.090 ± 0.052 pM) in 2010 and from 0.066 to 0.094 pM
(mean 0.079 ± 0.010 pM) in 2011. In general, DMMHg
was 19% to 78% of UMMHg (mean 45 ± 18%) and spatial
and temporal variations in DMMHg followed those of
UMMHg. There was a distinct spatial variation for particulate MMHg (PMMHg), which was significantly lower
(p < 0.05) in the saline zone compared to the upper- and
lower-river zones.
4. DISCUSSION
4.1. Variation of sediment Hg

In the Tien River, TOC was a major factor governing
lateral and temporal distribution of sediment Hg (Table 3,
Fig. 4). A similar correlation was found in other rivers and
estuaries (Hammerschmidt and Fitzgerald, 2004; Heyes
et al., 2006; Hollweg et al., 2009); this may be attributed
to strong affinity between Hg and thiolic binding sites on
TOC (Ravichandran, 2004; Skyllberg et al., 2007). No linear correlation was found between TOC and sediment Hg
for the Saigon River, with a relatively narrow range of
TOC. Point source location might be more important for
the Saigon River, since relatively high sediment Hg levels
were found from the runoff sites of Dau Tieng Reservoir
and Ho Chi Minh City. Increased metal concentrations
(e.g., Cd, Cr, Cu, and Zn) in urban parts of the Saigon River have been reported previously, and untreated urban and
industrial wastewaters were identified as major sources of
metal pollution (Thuy et al., 2007).
4.2. Variation of sediment MMHg

Fig. 3. Unfiltered Hg (UHg), dissolved Hg (DHg), and particulate
Hg (PHg) with distance to the coast in 2010 (a) and 2011 (b), and
unfiltered MMHg (UMMHg), dissolved MMHg (DMMHg), and
particulate MMHg (PMMHg) with distance to the coast in 2010 (a)
and 2011 (b) in Tien River.

Francisco Bay (0.050–2.3 pM; Conaway et al., 2003). Except for the steep decrease in the upper river, there was
no significant spatial variation for UMMHg across study
sites and there were no significant (p > 0.05) differences between 2010 and 2011. Dissolved MMHg (DMMHg) in the

The sediment MMHg in the Tien and Saigon Rivers
showed significant linear correlation with AVS (Table 3),
suggesting that biogeochemical processes that produce mineral FeS are associated with MMHg production. In 2010,

peak MMHg percentages were recorded at the estuarine
head of the Tien (site 10) and Saigon Rivers (site 4;
Fig. 5). Enhanced sulfate, associated with salinity intrusion,
appears to increase Hg(II) methylation rate and AVS (Hollweg et al., 2009). Relatively low MMHg percentage in the
saline zone, despite large AVS, could be related to the inhibition effect of dissolved sulfide (Benoit et al., 2001; Drott
et al., 2007). A number of studies report that dissolved sulfide shows strong inhibition of Hg(II) methylation rate in
coastal sediments (Conaway et al., 2003; Heyes et al.,
2006; Sunderland et al., 2006). Interestingly, in the present
study, there is a significant positive linear correlation between TN and MMHg in both rivers (Table 3). Assuming
that TN is more inorganic (e.g., ammonium) than organic,
based on the lack of correlation between TOC and TN, a
sediment redox condition might explain high MMHg. It
is commonly shown that reducing sediment provides favorable conditions for microbial Hg(II) methylation (Sunderland et al., 2006). Applying a multiple linear regression
for MMHg, it was found that AVS explained 41% of variability in MMHg (p < 0.05, n = 17) and that TN explained
an additional 15% of variability in MMHg ([MMHg


386

S. Noh et al. / Geochimica et Cosmochimica Acta 106 (2013) 379–390

Table 3
Partial correlation matrix for Hg, MMHg, %MMHg against
salinity, TOC, TN, AVS, Hg, and MMHg in sediment samples of
the Tien River (a) and Saigon River (b).
Hg
(nmol gÀ1)

MMHg
(pmol gÀ1)

%MMHg

Salinity (ppt)

À0.20
0.38
21

À0.13
0.58
21

0.083
0.72
21

TOC (%)

0.81
<0.0001
19

0.40
0.086
19

À0.17
0.47
19

TN (%)

À0.34
0.15
19

0.57
0.010
19

0.59
0.0081
19

À0.22
0.37
19

0.65
0.002
19

0.88
<0.0001
19

0.18
0.44
21

À0.42
0.056
21

(a) Tien River

À1

AVS (lmol g )

Hg (nmol gÀ1)

MMHg (pmol gÀ1)

0.73
0.0001
21

(b) Saigon River
Salinity (ppt)

0.38
0.35
8

0.44
0.28
8

0.23
0.59
8

TOC (%)

À0.50
0.20
8

0.29
0.48
8

0.40
0.32
8

TN (%)

0.53
0.17
8

0.86
0.0066
8

0.69
0.059

8

AVS (lmol gÀ1)

0.35
0.39
8

0.84
0.0085
8

0.66
0.077
8

0.24
0.57
8

À0.11
0.79
8

Hg (nmol gÀ1)

MMHg (pmol gÀ1)

0.93
0.00088

8

Listed from top to bottom: Pearson’s correlation coefficient, pvalue and sample number. Significant correlations are indicated in
bold type.

(pmol gÀ1)] = 0.19[AVS (lmol gÀ1)] + 7.45[TN (%)] À 1.7,
p < 0.05, n = 17). Besides AVS and sediment redox condition, TOC often plays a main role in controlling MMHg
(or %MMHg) in marine sediment (Conaway et al., 2003;
Hammerschmidt and Fitzgerald, 2004; Lambertsson and
Nilssons, 2006; Sunderland et al., 2006; Hollweg et al.,
2009). However, sediment TOC did not exhibit significant
correlation with MMHg in the Tien and Saigon Rivers (Ta-

Fig. 4. Correlation between Hg and total organic carbon (TOC) in
sediments of Tien and Saigon Rivers (linear regression includes
Tien data only).

ble 3), implying that organic content was not a limiting factor for Hg(II) methylation rate across our study sites.
In the Tien River, sediment %MMHg levels for 2011
were four to five times those for 2010, despite similar levels
of AVS (Table 2). The range of C/N was significantly lower
in 2011 (3.4–6.5) than in 2010 (9.6–15), reflecting larger
contributions of fresh and labile organic matter in 2011. Increased %MMHg (or Hg(II) methylation rate) influenced
by enrichment of fresh organic matter has been reported
from Venice Lagoon (Kim et al., 2011) and Hg-contaminated sites in Sweden (Drott et al., 2007). Enrichment of
fresh organic matter in 2011 sediment may be related to enhanced primary production (Table 2). Nutrients are the
main factor controlling phytoplankton growth in this region. The 2011 sampling campaign was indeed preceded
by a rainfall event lasting several days, which may have
caused eutrophication and excessive algal bloom.
4.3. Variation of Hg and MMHg in surface water

PHg and PMMHg content were not laterally uniform:
Hg- and MMHg-enriched particles transported by river
runoff appear to be diluted with Hg- and MMHg-depleted,
resuspended particles in the high turbidity zone
(log [SPM] > 1.9, Fig. 6). Indeed, the ranges of PHg and
PMMHg concentrations in the brackish zone were quite
similar to sediment Hg and MMHg concentrations, respectively (Table 2). Dilution of PHg and PMMHg in the estuarine high turbidity zone was opposite the findings reported
for the Seine Estuary, where matured particles (high C/N
ratio) in the high turbidity zone actually entrapped more
Hg and MMHg from the aqueous phase, resulting in increases in PHg and PMMHg in the estuarine high turbidity
zone (Laurier et al., 2003).
From Fig. 7a and b, it is evident that low POC content
in brackish particles explains PHg and PMMHg decreases
in the estuarine high turbidity zone. Hg data for 2011
showed relatively steady and low PHg levels, with no positive linear correlation with POC (Fig. 7a). The ratio of Chla to POC was used to differentiate suspended particles dom-


S. Noh et al. / Geochimica et Cosmochimica Acta 106 (2013) 379–390

387

Fig. 5. Distribution of avid-volatile sulfide (AVS) and %MMHg at
each sampling site for 2010 Tien (a) and 2010 Saigon (b) Rivers.
AVS data are missing in sites 11 and 12 for (a).
Fig. 7. Correlation between particulate Hg (PHg) and particulate
organic carbon (POC) (a) and between particulate MMHg
(PMMHg) and POC (b) in surface waters of Tien River. Linear
regression model for (a) includes 2010 data only.

Fig. 6. Correlation between particulate Hg (PHg) or particulate

MMHg (PMMHg) and suspended particulate matter (SPM) in
surface waters of Tien River.

This suggests that biodilution may be one reason for low
and steady PHg in 2011. A similar dilution effect for
PHg, associated with phytoplanktonic bloom, was reported
from the Seine Estuary (Laurier et al., 2003). For 2011, the
biodilution effect appears to be weak for MMHg, since linear regression was not distinct between 2010 and 2011
(Fig. 7b). This could be related to higher MMHg production in 2011, shown as sediment %MMHg in Table 2. Luengen and Flegal (2009) report that as the bloom decays,
MMHg in estuarine water significantly increases, likely
due to MMHg production from sediment and/or remineralization from decaying phytoplankton.
4.4. Partitioning of Hg and MMHg between solution and
particles

inated by labile organic matter (Chl-a/POC > 5 lg mgÀ1)
from those dominated by refractory organic matter (Cifuentes et al., 1988; Liu et al., 2007; Supporting Information,
Fig. S1). Labile organic particles cover a narrow range of
C/N, from 4.6 to 6.4, similar to the Redfield ratio. For
2010, it is notable that freshwater SPM is dominated by labile organics, while brackish SPM is dominated by refractory organics. For 2011, however, most particles in
freshwater and brackish zones were labile; this might be
associated with an excessive algal bloom in 2011 (Table 1).

The log-transformed particle–water partition coefficient,
Kd = [particulate Hg] (mol kgÀ1)/[dissolved Hg] (mol LÀ1),
of Hg averaged 5.1 ± 0.44 in 2010 and 4.9 ± 0.21 in 2011
in the Tien River (Fig. 8). These ranges were similar to
those found from other rivers (e.g., 4.8–5.7 in the Patuxent
River, Benoit et al., 1998; 4.5–6.5 in the St. Lawrence River,
Que´merais et al., 1998; and 2.8–6.6 for streams in Oregon
and Wisconsin, Brigham et al., 2009). The log Kd of MMHg

averaged 4.5 ± 0.77 in 2010 and 4.5 ± 0.50 in 2011 in the


388

S. Noh et al. / Geochimica et Cosmochimica Acta 106 (2013) 379–390

and log K0 = 13.6, Fitzgerald et al., 2007). There appears
to be more than enough L with sufficient affinity with Hg
to compete out Hg–chloro complexes, but this is not the
case for MMHg in the estuarine zone of the Tien River.
5. SUMMARY

Fig. 8. Kd (particle–water partition coefficient) of Hg, MMHg and
organic carbon (OC) as a function of suspended particulate matter
(SPM) concentration in Tien River.

Tien River. We found a negative linear relationship between log Kd of Hg (MMHg and OC) and log [SPM]; this
is known as a particle concentration effect (Benoit, 1995;
Fig. 8). The POC content in SPM may play a critical role
in this correlation, as we found less POC under high SPM
conditions. However, there should be additional factors
to explain the close negative correlation, since POC contents were relatively constant when log [SPM] > 1.9 (Supporting Information, Table S1). Increasing colloidal
organic matter along with increasing SPM may be associated with decreases in Kd of Hg, MMHg, and OC as a function of SPM (Honeyman and Santschi, 1991; Benoit and
Rozan, 1999; Lee et al., 2011).
The slope (À0.52) of the linear regression between log Kd
of Hg and log [SPM] was similar to that (À0.38) between
log Kd of OC and log [SPM], indicating the critical role of
organic matter in governing phase-partitioning of Hg (Benoit et al., 1998; Choe and Gill, 2003; Conaway et al.,
2003; Laurier et al., 2003; Sunderland et al., 2006). Interactions between organic matter and Hg, attributable to strong

Hg binding with reduced sulfur-containing functional
groups (e.g., thiol), appear to control particle–water partition of Hg. On the other hand, the slope (À0.97) of the linear regression between log Kd of MMHg and log [SPM] was
substantially lower than those for Hg and OC. We speculate that chloro-complexation of MMHg provides additional decreases in log Kd of MMHg in the estuarine zone
(log [SPM] > 1.9). It is known that interactions between
MMHg and organic ligand (L) are less significant compared to those between Hg and L (Zhong and Wang,
2009). According to the Hg–Cl–L ligand complexation
model, HgL complex dominates dissolved Hg pool under
the Tien’s estuarine conditions ([Cl] = 0.02–0.3 M,
[DOC] = 180–360 lM), assuming L/DOC = (5–50) Â 10À6
and 22 < log K0 < 25, Fitzgerald et al., 2007). On the contrary, the CH3Hg–Cl–L ligand complexation model predicts that CH3HgL competes with CH3HgCl under our
estuarine conditions, assuming L/DOC = (5–50) Â 10À6

The Mekong River Delta, located in Vietnam, is a flat,
low-lying area of highly complex rivers, channels, and flood
plains. Although there are more than 1000 man-made canals in the Mekong Delta for inhibition of saltwater, transport, and land reclamation, saline intrusion still occurs and
causes severe human suffering during dry seasons (Hoa
et al., 2007). In surface sediment, enhanced %MMHg was
found at the estuarine head, along with increased AVS,
emphasizing the importance of sulfate availability. In surface water, UHg concentrations were greater in the estuarine high turbidity zone compared to the upper and lower
rivers, due to enhanced particle load. Fractions of particulate Hg appear to be remobilized in the estuarine high turbidity zone, since DHg also increased in the high turbidity
zone. On the contrary, UMMHg and DMMHg did not increase in the estuarine high turbidity zone compared to the
upper and lower rivers, likely due to large decreases in
PMMHg. Although the high turbidity zone appears to have
elevated microbial activity, which manifests as low ORP,
sulfide may play a critical role in constraining active Hg(II)
methylation. Regarding particle–water distribution of Hg
and MMHg, we found increased solubility of Hg and
MMHg in the estuarine zone, with a negative linear relationship between log Kd of Hg (MMHg) and log [SPM]. Between 2010 and 2011, sediment and water conditions were
highly variable in terms of Hg and organic quality. Hg in
suspended particles was depleted in the bloom year

(2011), highlighting the importance of the biodilution effect.
In contrast, sediment %MMHg increased in the bloom
year, perhaps due to enrichment of labile fresh organic matter in sediment, and subsequent enhancement of microbial
activity. Taken as a whole, Hg speciation and partitioning
in the lower Mekong River Basin were strongly influenced
by salinity intrusion, and major biogeochemical factors
affecting Hg behaviors were particle loads, biological productivity, and concentrations of sulfate, chloride and organic matter.
ACKNOWLEDGEMENTS
We are grateful for the support of the Hanyang Research and
Industry Cluster at Hanyang University. This study was supported
by the Ministry of Science and Technology, Korea, through the
Institute of Science and Technology for Sustainability (UNU &
GIST Joint Program); and by the Ministry of Land, Transport,
and Maritime Affairs through “Impacts of ocean acidification on
the bioaccumulation and release of mercury by microbes”.

APPENDIX A. SUPPLEMENTARY DATA
Supplementary data associated with this article can be
found, in the online version, at />j.gca.2012.12.018.


S. Noh et al. / Geochimica et Cosmochimica Acta 106 (2013) 379–390
REFERENCES
Balcom P. H., Hammerschmidt C. R., Fitzgerald W. F., Lamborg
C. H. and O’Connor J. S. (2008) Seasonal distributions and
cycling of mercury and methylmercury in the waters of New
York/New Jersey Harbor Estuary. Mar. Chem. 109, 1–17.
Baran E., Jantunen T. and Chong C. K. (2007) Values of Inland
Fisheries in the Mekong River Basin. Worldfish Center.
Benoit G. (1995) Evidence of the particle concentration effect for

lead and other metals in fresh waters based on ultraclean
technique analyses. Geochim. Cosmochim. Acta 59, 2677–2687.
Benoit G. and Rozan T. F. (1999) The influence of size distribution
on the particle concentration effect and trace metal partitioning
in rivers. Geochim. Cosmochim. Acta 63, 113–127.
Benoit J. M., Gilmour C. C., Mason R. P., Riedel G. S. and Riedel
G. F. (1998) Behavior of mercury in the Patuxent River estuary.
Biogeochemistry 40, 249–265.
Benoit J. M., Gilmour C. C. and Mason R. P. (2001) The influence
of sulfide on solid-phase mercury bioavailability for methylation by pure cultures of Desulfobulbus propionicus (1pr3).
Environ. Sci. Technol. 35, 127–132.
Brigham M. E., Wentz D. A., Aiken G. R. and Krabbenhoft D. P.
(2009) Mercury cycling in stream ecosystems. 1. Water column
chemistry and transport. Environ. Sci. Technol. 43, 2720–2725.
Cana´rio J., Vale C. and Nogueira M. (2008) The pathway of
mercury in contaminated waters determined by association with
organic carbon (Tagus Estuary Portugal). Appl. Geochem. 23,
519–528.
Cenci R. M. and Martin J. M. (2004) Concentration and fate of
trace metals in Mekong River Delta. Sci. Total Environ. 332,
167–182.
Chen C., Amirbahman A., Fisher N., Harding G., Lamborg C.,
Nacci D. and Taylor D. (2008) Methylmercury in marine
ecosystems: spatial patterns and processes of production
bioaccumulation and biomagnification. EcoHealth 5, 399–408.
Chiffoleau J.-F., Cossa D., Auger D. and Truquet I. (1994) Trace
metal distribution partition and fluxes in the Seine estuary
(France) in low discharge regime. Mar. Chem. 47, 145–158.
Choe K. Y. and Gill G. A. (2003) Distribution of particulate
colloidal and dissolved mercury in San Francisco Bay estuary.

2. Monomethyl mercury. Limnol. Oceanogr. 48, 1547–1556.
Choe K. Y., Gill G. A. and Lehman R. (2003) Distribution of
particulate colloidal and dissolved mercury in San Francisco
Bay estuary. 1. Total mercury. Limnol. Oceanogr. 48, 1535–
1546.
Cifuentes L. A., Sharp J. H. and Fogel M. L. (1988) Stable carbon
and nitrogen isotope biogeochemistry in the Delaware estuary.
Limnol. Oceanogr. 33, 1102–1115.
Conaway C. H., Squire S., Mason R. P. and Flegal A. R. (2003)
Mercury speciation in the San Francisco Bay estuary. Mar.
Chem. 80, 199–225.
Covelli S., Piani R., Kotnik J., Horvat M., Faganeli J. and
Brambati A. (2006) Behaviour of Hg species in a microtidal
deltaic system: the Isonzo River mouth (northern Adriatic Sea).
Sci. Total Environ. 368, 210–223.
Dai A. and Trenberth K. E. (2002) Estimates of freshwater
discharge from continents: latitudinal and seasonal variations.
J. Hydrometeorol. 3, 660–687.
Drott A., Lambertsson L., Bjo¨rn E. and Skyllberg U. (2007)
Importance of dissolved neutral mercury sulfides for methyl
mercury production in contaminated sediments. Environ. Sci.
Technol. 41, 2270–2276.
Edwin D. O. (2009) Water quality of the lower Mekong River. In
The Mekong: Biophysical Environment of an International River
Basin (ed. I. C. Campbell). Academic Press, pp. 297–320.

389

Environmental Protection Agency (1996) Method 9030B; Acidsoluble and Acid-insoluble Sulfides: Distillation. EPA-SW-846Ch-5.
Environmental Protection Agency (2001) Method 1630; Methyl

Mercury in Water by Distillation Aqueous Ethylation Purge
and Trap and CVAFS. EPA-821-R-01-202.
Environmental Protection Agency (2001) Appendix to Method
1631; Total Mercury in Tissue, Sludge, Sediment, and Soil by
Acid Digestion and BrCl Oxidation. EPA-821-R-01-013.
Environmental Protection Agency (2002) Method 1631 Revision E;
Mercury in Water by Oxidation Purge and Trap and Col Vapor
Atomic Fluorescence Spectrometry. EPA-821-R-02-019.
Fitzgerald W. F., Lamborg C. H. and Hammerschmidt C. R.
(2007) Marine biogeochemical cycling of mercury. Chem. Rev.
107, 641–662.
Gantner N., Power M., Babaluk J. A., Reist J. D., Ko¨ck G.,
Lockhart L. W., Solomon K. R. and Muir D. C. (2009)
Temporal trends of mercury, cesium, potassium, selenium, and
thallium in arctic char (Salvelinus alpinus) from Lake Hazen
Nunavut Canada: effects of trophic position size and age.
Environ. Toxicol. Chem. 28, 254–263.
George B. A., Malano H. M., Khac Tri V. O. and Turral H. (2004)
Using modelling to improve operational performance in the Cu
Chi irrigation system Vietnam. Irrigat. Drainage 53, 237–249.
Hammerschmidt C. R. and Fitzgerald W. F. (2004) Geochemical
controls on the production and distribution of methylmercury
in near-shore marine sediments. Environ. Sci. Technol. 38,
1487–1495.
Hammerschmidt C. R. and Fitzgerald W. F. (2006) Methylmercury
cycling in sediments on the continental shelf of southern New
England. Geochim. Cosmochim. Acta 70, 918–930.
Hammerschmidt C. R., Fitzgerald W. F., Lamborg C. H., Balcom
P. H. and Visscher P. T. (2004) Biogeochemistry of methylmercury in sediments of Long Island Sound. Mar. Chem. 90,
31–52.

Han S., Obraztsova A., Pretto P., Choe K. Y., Gieskes J., Deheyn
D. D. and Tebo B. M. (2007) Biogeochemical factors affecting
mercury methylation in sediments of the Venice Lagoon Italy.
Environ. Toxicol. Chem. 26, 655–663.
Heyes A., Mason R. P., Kim E. H. and Sunderland E. (2006)
Mercury methylation in estuaries: insights from using measuring rates using stable mercury isotopes. Mar. Chem. 102, 134–
147.
Hoa L. T. V., Nhan N. H., Wolanski E., Cong T. T. and Shigeko
H. (2007) The combined impact on the flooding in Vietnam’s
Mekong River delta of local man-made structures sea level rise
and dams upstream in the river catchment. Estuar. Coast. Shelf
Sci. 71, 110–116.
Hollweg T. A., Gilmour C. C. and Mason R. P. (2009) Methylmercury production in sediments of Chesapeake Bay and the
mid-Atlantic continental margin. Mar. Chem. 114, 86–101.
Honeyman B. D. and Santschi P. H. (1991) Coupling adsorption
and particle aggregation: laboratory studies of “colloidal
pumping” using 59Fe-labeled hematite. Environ. Sci. Technol.
25, 1739–1747.
Kim E. H., Mason R. P., Porter E. T. and Soulen H. L. (2006) The
impact of resuspension on sediment mercury dynamics and
methylmercury production and fate: a mesocosm study. Mar.
Chem. 102, 300–315.
Kim M., Han S., Gieskes J. and Deheyn D. D. (2011) Importance
of organic matter lability for monomethylmercury production
in sulfate-rich marine sediments. Sci. Total Environ. 409, 778–
784.
Lambert M. S., Mariam T. T. and Susan F. H. (2010) Saigon River.
Kniga po Trebovaniyu Press.

390

S. Noh et al. / Geochimica et Cosmochimica Acta 106 (2013) 379–390

Lambertsson L. and Nilssons M. (2006) Organic material: the
primary control on mercury methylation and ambient methyl
mercury concentrations in estuarine sediments. Environ. Sci.
Technol. 40, 1822–1829.
Laurier F. J. G., Cossa D., Gonzalez J. L., Breviere E. and Sarazin
G. (2003) Mercury transformations and exchanges in a high
turbidity estuary: the role of organic matter and amorphous
oxyhydroxides. Geochim. Cosmochim. Acta 67, 3329–3345.
Lee S., Han S. and Gill G. A. (2011) Estuarine mixing behavior of
colloidal organic carbon and colloidal mercury in Galveston
Bay Texas. J. Environ. Monitor. 13, 1703–1708.
Liu K. K., Kao S. J., Wen L. S. and Chen K. L. (2007) Carbon and
nitrogen isotopic compositions of particulate organic matter
and biogeochemical processes in the eutrophic Danshuei
Estuary in northern Taiwan. Sci. Total Environ. 382, 103–120.
Liu J., Feng X., Zhu W., Zhang X. and Yin R. (2012) Spatial
distribution and speciation of mercury and methyl mercury in
the surface water of East River (Dongjiang) tributary of Pearl
River Delta South China. Environ. Sci. Pollut. R. 19, 105–112.
Luengen A. C. and Flegal A. R. (2009) Role of phytoplankton in
mercury cycling in the San Francisco Bay estuary. Limnol.
Oceanogr. 54, 23–40.
Minh N. H., Minh T. B., Iwata H., Kajiwara N., Kunisue T.,
Takahashi S., Viet P. H., Tuyen B. C. and Tanabe S. (2007)
Persistent organic pollutants in sediments from Sai Gon-Dong
Nai River basin Vietnam: levels and temporal trends. Arch.

Environ. Contam. Toxicol. 52, 458–465.
Morse J., Thomson H. and Finneran D. (2007) Factors controlling
sulfide geochemistry in sub-tropical estuarine and bay sediments. Aquat. Geochem. 13, 143–156.
Osborne M. (2000) The Mekong: Turbulent Past Uncertain Future.
Grove Press.
Pan J. F. and Wang W. X. (2004) Uptake of Hg(II) and
methylmercury by the green mussel Perna viridis under different
organic carbon conditions. Mar. Ecol. Prog. Ser. 276, 125–136.
Peckenham J. M., Kahl J. S. and Mower B. (2003) Background
mercury concentrations in river water in Maine U.S.A. Environ.
Monitor. Asses. 89, 129–152.
Que´merais B., Cossa D., Rondeau B., Pham T. T. and Fortin B. (1998)
Mercury distribution in relation to iron and manganese in the
waters of the St. Lawrence river. Sci. Total Environ. 213, 193–201.
Ramalhosa E., Pereira E., Vale C., Va´lega M., Monterroso P. and
Duarte A. C. (2005) Mercury distribution in Douro estuary
(Portugal). Mar. Pollut. Bull. 50, 1218–1222.
Ravichandran M. (2004) Interactions between mercury and
dissolved organic matter – a review. Chemosphere 55, 319–331.
Roth D. A., Taylor H. E., Domagalski J., Dileanis P., Peart D. B.,
Antweiler R. C. and Alpers C. N. (2001) Distribution of inorganic
mercury in Sacramento River water and suspended colloidal
sediment material. Arch. Environ. Contam. Toxicol. 40, 161–172.

Schuster P. F., Shanley J. B., Marvin-Dipasquale M., Reddy M.
M., Aiken G. R., Roth D. A., Taylor H. E., Krabbenhoft D. P.
and DeWild J. F. (2008) Mercury and organic carbon dynamics
during runoff episodes from a northeastern USA watershed.
Water, Air, Soil Pollut. 187, 89–108.
Skyllberg U., Drott A., Lambertsson L., Bjo¨rn E., Karlsson T.,

˚ . and Holmstro¨m H. (2007) Net
Johnson T., Heinemo S. A
methylmercury production as a basis for improved risk assessment of mercury-contaminated sediments. Ambio 36, 437–442.
Stordal M. C., Gill G. A., Wen L. S. and Santschi P. H. (1996)
Mercury phase speciation in the surface waters of three Texas
estuaries: importance of colloidal forms. Limnol. Oceanogr. 41,
52–61.
Sunderland E. M., Gobas F. A. P. C., Branfireun B. A. and Heyes
A. (2006) Environmental controls on the speciation and
distribution of mercury in coastal sediments. Mar. Chem. 102,
111–123.
Thuy H. T. T., Vy N. N. H. and Loan T. T. C. (2007)
Anthropogenic input of selected heavy metals (Cu, Cr, Pb,
Zn, and Cd) in the aquatic sediments of Hochiminh city
Vietnam. Water, Air, Soil Pollut. 182, 73–81.
Turner A., Millward G. E. and Le Roux S. M. (2004) Significance
of oxides and particulate organic matter in controlling trace
metal partitioning in a contaminated estuary. Mar. Chem. 88,
179–192.
Vo P. L. (2007) Urbanization and water management in Ho Chi
Minh City Vietnam-issues challenges and perspectives. GeoJournal 70, 75–89.
Wolanski E., Huan N. N., Dao L. T., Nhan N. H. and Thuy N. N.
(1996) Fine-sediment dynamics in the Mekong River Estuary
Vietnam. Estuar. Coast. Shelf Sci. 43, 565–582.
Wolanski E., Nhan N. H. and Spagnol S. (1998) Sediment
dynamics during low flow conditions in the Mekong river
estuary Vietnam. J. Coast. Res. 14, 472–482.
Xue Z., Liu J. P. and Ge Q. (2011) Changes in hydrology and
sediment delivery of the Mekong River in the last 50 years:
connection to damming monsoon and ENSO. Earth Surf. Proc.

Land. 36, 296–308.
Xue Z., Liu J. P., DeMaster D., Van Nguyen L. and Ta T. K. O.
(2010) Late Holocene evolution of the Mekong Subaqueous
Delta Southern Vietnam. Mar. Geol. 269, 46–60.
Zhang L. and Wong M. H. (2007) Environmental mercury
contamination in China: sources and impacts. Environ. Int.
33, 108–121.
Zhong H. and Wang W. X. (2009) Inorganic mercury binding with
different sulfur species in anoxic sediments and their gut juice
extractions. Environ. Toxicol. Chem. 28, 1851–1857.
Associate editor: Christopher Kim