DSpace at VNU: Heavy metal contamination of agricultural soils around a chromite mine in Vietnam

Soil Science and Plant Nutrition (2010) 56, 344–356

doi: 10.1111/j.1747-0765.2010.00451.x

ORIGINAL ARTICLE

Heavy metal contamination of agricultural soils around a chromite
mine in Vietnam
Chu N. KIEN1, Nguyen V. NOI2, Le T. SON2, Ha M. NGOC2, Sota TANAKA3,
Takuro NISHINA4 and K
oz
o IWASAKI4
1

United Graduate School of Agricultural Sciences, Ehime University, Ehime 790-8566, 2Faculty of Chemistry, Hanoi University of
Science, Hanoi, Vietnam, 3Graduate School of Kuroshio Science and 4Faculty of Agriculture, Kochi University, Kochi 783-8502, Japan

Abstract
In Vietnam, the Co Dinh mine is the largest chromite mine in the country. Mining, ore dressing and disposal of
the tailings provide obvious sources of heavy metal contamination in the mine area. The present study examined
the influence of chromite mining activities on the adjacent lowland paddy field by investigating heavy metal and
As levels in the mine tailings, sediments, paddy soils and water. At paddy fields located near the mine tailings,
the total contents of Cr, Co and Ni were 5,750, 375 and 5,590 mg kg)1, and the contents of their water-extractable form were 12.7, 1.16 and 32.3 mg kg)1, respectively. These results revealed severe contamination of lowland paddy soils with Cr, Co and Ni as a result of mining activity, suggesting serious health hazards through
agricultural products, including livestock in this area. The principal source of the pollution was sediment inflow
owing to the collapse of the dike, which was poorly constructed by heaping up soil. Moreover, water flowing
out from the mining area was also polluted with Cr and Ni (15.0–41.0 and 20.0–135 lg L)1, respectively). This
might raise another problem of heavy metal pollution of watercourses in the area, indicating the need for further
investigation and monitoring of fluctuations of water quality with seasonal changes.
Key words:

contamination, heavy metal, mine, soil, Vietnam.

INTRODUCTION
Mining activities produce large quantities of waste materials, such as waste rock, tailings and slag, leading to metal
contamination of the environment (Adriano 2001; Chopin and Alloway 2007; Jung 2001). Elevated levels of
toxic metals are often reported in agricultural soils, food
crops and stream systems as a result of the discharge and
dispersion of mine wastes into the environment (Jung
2001; Lee 2006; McGowen and Basta 2001). A number
of studies have investigated the spatial distribution and
behavior of heavy metals in and around mining areas to
assess the potential health risks and environmental hazards caused by polluted agricultural products. For example, enrichment of Cr and Ni (86–358 and 21.2–
126 mg kg)1, respectively) as a result of mine tailings was
reported for surface soils taken from the Almade´n mining
Correspondence: C. N. KIEN, United Graduate School of Agricultural Sciences, Ehime University, 3-5-7, Tarumi, Matsuyama,
Ehime 790-8566, Japan. Email:
Received 25 September 2009.
Accepted for publication 20 December 2009.

district in Spain (Bueno et al. 2009). Elevated levels of Cr
and Ni (182–1,029 mg Cr kg)1 and 15–432 mg Ni kg)1)
were also found in soils collected from mining areas in
southern Togo (Gnandi and Tobschall 2002). Bi et al.
(2006) reported Cr contamination of agricultural soils
(71–240 mg kg)1), resulting from zinc smelting activities
in Hezhang County, western Guizhou, China. High levels
of heavy metals in paddy fields nearby the Daduk Au–Ag–
Pb–Zn mining area of Korea were caused by dispersion of
the metals from the tailings by sedimentation and watercourses (Lee et al. 2001). In addition, a number of studies
have investigated the vertical distribution of heavy metals
in soil profiles in and round mining areas, and changes in

the geochemical processes throughout the profile (Adriano
2001; Johnson et al. 2000; Kim and Jung 2004; Otero
et al. 2000). According to these studies, it appears that the
profile distribution patterns of each heavy metal are site
specific, with no consistent distribution pattern correlating
with soil depth.
The toxicity and bioavailability of heavy metals in soils
are influenced by the metal’s mobility and reactivity with
various environmental factors. Therefore, information
about metal speciation as well as its total content is necesÓ 2010 Japanese Society of Soil Science and Plant Nutrition


Soil contamination by a chromite mine in Vietnam

sary for the assessment of heavy metal toxicity and bioavailability. The proportion of a metal that is mobile and
bioavailable will provide practical information for evaluating its potential environmental risks. Kien et al. (2009)
studied and reported the form and horizontal distribution
of Cu and As in paddy soils and watercourses resulting
from the exploitation of tin and tungsten ores in Daitu district, in the northern part of Vietnam. However, in Vietnam, few studies on the form and distribution of heavy
metals and metalloids have been carried out for agricultural soils affected by mining activities.
In Vietnam, a wide variety of minerals have been found
to contain various elements (e.g. antimony, chromite, copper, tin, tungsten), and a lot of metalliferrous mines have
been established over the country. Some are currently
being mined, whereas others have been abandoned. Many
of these mines are located in mountainous areas or in the
upper reaches of lowland streams, where various types of
crops are cultivated, including lowland rice, which is the
major crop. Although dikes are usually constructed
around mine areas to prevent the release of tailings, waste
water and solid waste into the surrounding environment,

frequent occurrences of flooding during the rainy season
have caused some of these dikes to collapse and not
function properly, resulting in heavy metal pollution in
lower streams and farmland areas. Therefore, it is impera-

Figure 1 Location of the sampling sites.
Ó 2010 Japanese Society of Soil Science and Plant Nutrition

345

tive that we accumulate more data related to heavy metal
contamination around mining sites where various types of
metals are extracted.
In the present study, we assessed the influence of heavy
metal contamination on lowland paddy fields in the
downstream areas of the Co Dinh chromite mine and clarified possible pathways that the contaminants might take.
The chemical forms of the contaminants were also determined to evaluate their mobility and any potential risks to
the surrounding environment.

MATERIALS AND METHODS
Co Dinh chromite mine
Mining activity at the Co Dinh chromite mine (19o43¢N,
105o36¢E; Fig. 1) was initiated in the early 20th century
(1930) with ore extraction from the ground surface, followed by open-pit mining. Large-scale and intensive mining activities commenced in the 1990s. Recently, the Co
Dinh Chromite Mining Company temporarily closed the
mine with the aim of re-construction and installation of
new technology. However, local people living in the vicinity of the mine continue small-scale activities using small
adits and open pits.
The chromite concentrate produced from this mine
contains 46% Cr2O3 and <27% Fe2O3, 5% SiO2 and

346 C. N. Kien et al.

0.4% H2O; the estimated ore reserve of the mine is
approximately 20.8 metric tons of Cr2O3 (Wu 2002,
2004). Annual production of the concentrate reached
80,000 tons per year.
Surface and open-pit mining, followed by on-site ore
processing have left a huge number of waste rock piles,
spoil heaps and tailing ponds within approximately 2 km2
of the mining area. These mine waste residues are transported by small streams, eventually reaching and being
deposited into a large lake (approximately 0.5 km2); the
lake was formed by the intensive mining activities (Fig. 1).
A number of local people constructed a dike at the border
of the mining site by heaping up soil to prevent pollutant
inflow into the adjacent agricultural lands. However, the
dike sometimes collapses as a result of flooding during the
rainy season, resulting in the inflow of sediments, including mine wastes, into the adjacent drainage canals and
paddy field areas.

Study sites and sampling
Field surveys and sampling were conducted in November
2006 and November 2007 at study sites selected within
the Co Dinh chromite mine and in the adjacent lowland
paddy fields (Fig. 1; Table 1). The climate is tropical monsoon with a mean annual precipitation and temperature
of 1,600–2,000 mm and 23–24°C, respectively (Thanh
Hoa Department of Culture, Sport and Tourism 2009).

At the time of the field surveys and sampling (November

2006 and November 2007), rice plants had just been harvested and the paddy fields were submerged for the next
rice cultivation.
Three representative sites were examined to characterize the soil properties of the study area and to investigate
the vertical distribution of heavy metals and As with soil
depth: one site was located in the mining area (Sp1) and
two sites were located in the areas used for paddy fields,
severely affected (Sp2) and non-affected (Sp3) by mine
wastes. At each site, a soil pit was prepared and its profile
was described. Soil samples were taken from each horizon
of the soil profiles, placed into a plastic bag and stored at
4°C until analysis.
To assess the influence of heavy metals and As from the
mining areas and to clarify the pathways of contamination, samples of mine tailings, sediment, paddy soil and
water were collected from established sites at various distances from the mining area. Tailings was taken from a
depth of 0–25 cm at the mine tailings site (designated as
MT). Stream sediments were sampled from the bottom of
streams running through the mining area (SD1–4 and
SD6). In addition, one sediment sample was taken from
the lake (SD5; Fig. 1). In total, eight paddy fields were
selected as study sites: six paddy fields (P1–6) were
assumed to be affected by sediment inflow as a result of
the dike collapse, and two paddy fields (P7–8) were

Table 1 List of the sampling sites

Site

Distance and direction
from tailings (km)

Location

Soil pits
Sp1
Located on a mountain slope of a natural laurel forest
Sp2
A paddy field located immediately outside of the dike
Sp3
A paddy field located far from the dike (between sites P7 and P8)
Mine tailings and sediments
MT
Mine tailings
SD1
Upstream running through the mine tailings
SD2
Downstream of site SD1
SD3
Downstream of site SD2
SD4
Downstream of site SD3 and adjacent to the reservoir lake
SD5
Center of the lake
SD6
Located at the drainage channel running from
the reservoir lake and flowing into paddy fields
Paddy soils
P1
A paddy field located inside of the dike,
the transition ⁄ buffering area between the lake and the dike
P2

A paddy field located immediately outside of the dike
P3
A paddy field located outside of the dike
P4
A paddy field located outside of the dike
P5
A paddy field located outside of the dike
P6
A paddy field located outside of the dike
P7
A paddy field located far from the mining sites
P8
A paddy field located far from the mining sites

Sample type

0.00 NE
2.70 NE
5.30 NE

Mine soil
Paddy soil
Paddy soil

0.00 NE
0.30 NE
0.50 NE
0.80 NE
1.50 NE
2.00 NE

2.40 NE

Tailings
Stream sediment
Stream sediment
Stream sediment
Stream sediment
Lake sediment
Stream sediment

2.50 NE

Paddy soil

2.60 NE
2.80 NE
3.00 NE
3.50 NE
4.50 NE
5.00 NE
6.00 NE

Paddy soil
Paddy soil
Paddy soil
Paddy soil
Paddy soil
Paddy soil
Paddy soil

Ó 2010 Japanese Society of Soil Science and Plant Nutrition


Soil contamination by a chromite mine in Vietnam

assumed to be unaffected because the transportation road
formed an effective barrier. At each paddy field, surface
(0–5 cm) and subsurface (20–25 cm) soils were collected.
In addition to sampling the tailings, sediments and soils,
water samples were collected from each site, including
stream water near the MT, SD1 and SD3–4 sites, lake
water at SD5 site, and standing water in the paddy fields
(P1–2 and P4–8).

Sample analyses
Soil and sediment analyses
General physicochemical properties. The soil and sediment samples were air-dried at room temperature and
crushed to pass through a 2-mm mesh sieve. The particle
size distribution was determined using the pipette method
(Gee and Bauder 1986). The clay mineral composition
was identified by X-ray diffraction (XRD) analysis using
CuKa radiation (XD-D1w; Shimadzu, Kyoto, Japan). The
electrical conductivity (EC) and pH (H2O) values were
determined using a platinum electrode and a glass electrode, respectively, at 1:5 (w ⁄ v) ratio of soil ⁄ sediment : water. The pH and redox potential (Eh) of the
soils taken from the three soil profiles and from some
selected sediment samples were determined in situ.
Exchangeable (Ex-) cations (K, Na, Ca and Mg) were
extracted with 1 mol L)1 ammonium acetate at pH 7.0,
and the contents were determined using an atomic absorption spectrometer ([AAS] AA-6800; Shimadzu). After
removing excess NH4+, the sample was extracted with

100 g L)1 NaCl solution and the supernatant was used to
determine the cation exchange capacity (CEC) using the
Kjeldahl distillation and titration method (Rhoades
1982). The contents of total carbon (TC) and total nitrogen (TN) were analyzed on an NC analyzer (Sumigraph
NC-80; Sumitomo Chemical, Osaka, Japan).
Total heavy metals and As. After the samples were
digested in a mixture of HNO3 and HF (9:1) and heated
in a microwave (Multiwave, Perkin-Elmer, Yokohama,
Japan), the total concentration of the heavy metals (Cr,
Co, Cu, Ni and Pb) was determined using AAS. The accuracy of the method was assessed using certificated reference soils (JSO-1, provided by the National Institute of
Advanced Industrial Science and Technology, Tsukuba,
Japan) and marine sediment (NIES No.12, provided by
the National Institute for Environmental Studies, Tsukuba, Japan). The recoveries of Co, Cr, Cu, Ni and Pb
were in the ranges 92.6–117, 96.7–101, 96.6–101, 96.2–
104 and 92.9–96.2%, respectively.
For the analysis of total As content in the soils and sediments, samples were digested in a mixture of HClO4–
HNO3–HF (2:3:5) with the addition of 20 g L)1 KMnO4
in a teflon vessel at 100°C. The concentration of As in the
acid digest was determined using an inductively coupled
Ó 2010 Japanese Society of Soil Science and Plant Nutrition

347

plasma-atomic emission spectrometer ([ICP-AES] ICPS1000IV; Shimadzu) equipped with a hydride vapor generator (HVG-1; Shimadzu). The standard reference materials JSO-1 and JSO-2 (provided by the National Institute
of Advanced Industrial Science and Technology, Tsukuba,
Japan) were used to verify the accuracy of the As determination. The recovery rates of As were within 90–95%.
Water-extractable forms of the heavy metals and
As. Water-extractable forms of the heavy metals and As
were examined using fresh (moist) soil and sediment samples. The procedure for this extraction was adapted from
a previous report by Kim et al. (2003). In brief, a 10-g

portion of each sample and 50 mL of deionized water
were placed into a 100 mL plastic vessel. All samples were
shaken using a platform shaker at room temperature for
1 h. The suspended mixtures were filtered through
0.45 lm cellulose membrane filters. The concentrations
of the heavy metals and As in the leachates were determined by AAS and ICP-AES, respectively.
Extraction of hexavalent chromium. Extraction and
determination of total hexavalent chromium (Cr(VI)) in
the soils and sediments were conducted based on the EPA
Method 3060A ⁄ 7196A with some modifications. The
protocol applied here was described in detail by James
et al. (1995). In brief, 2.5 g of homogenized fresh (moist)
soil or sediment sample was placed into a 250 mL beaker.
Fifty milliliters of a solution of 0.28 mol L)1 Na2CO3 in
0.5 mol L)1 NaOH (pH 12) was added and mixed
(unheated) for approximately 10 min. The beakers were
then transferred to a preheated 150°C hot plate and maintained at 90–95°C for 60 min with continuous stirring.
Then, the solution was cooled and replenished with distilled water to the initial volume. After centrifugation for
phase separation (20 min, 2000 g), the supernatant was
filtered through a 0.45-lm cellulose membrane filter. The
solution was then adjusted to a pH of 7.5 ± 0.5 with
5.0 mol L)1 HNO3 solution and topped up to 100 mL
with distilled water. The sample digests were then analyzed using the colorimetric method. One milliliter of
diphenylcarbazide solution was added to 80 mL of digest
in 100 mL Erlenmeyer flasks. A 1.8 mol L)1 H2SO4 solution was added to the digests until the solution reached a
pH value between 1.6 and 2.2, then it was topped up to
100 mL with distilled water. After 10 min, the absorbance at 540 nm was measured using a spectrophotometer (V-360 BIO; Jasco, Tokyo, Japan) and the Cr(VI)
concentration was determined against standard solutions
ranging from 0.05 to 2 mg L)1.

Water analysis
Water samples were filtered through a 0.45-lm membrane filter and divided into two portions. One portion
was acidified with 0.03 mol L)1 HNO3 for analysis of
heavy metal and As concentrations, and the other portion


348 C. N. Kien et al.

was left unacidified for pH and EC measurements. The
water samples were then stored in a refrigerator at 4°C
until analysis. The total concentrations of heavy metals
(Co, Cr, Cu, Ni and Pb) and As were determined by AAS
and ICP-AES, respectively.

Sp2 soils was the highest throughout the profile among
the three sites. The XRD analysis revealed that the clay
minerals were composed mainly of montmorillonite, kaolin minerals, mica minerals and quartz for the Sp1 soils
and kaolin minerals, mica minerals, quartz and gibbsite
for the Sp3 soils. In the case of the Sp2 soils, both montmorillonite and gibbsite were detected in addition to the
other minerals, which were detected commonly in the Sp1
and Sp3 soils. Based on the USDA soil classification system, paddy soils at both the Sp2 and Sp3 sites were classified as Typic Endoaquepts, whereas the soil at the Sp1 site
was determined to be a Lithic Ustorthents (Soil Survey
Staff 2006).
As a whole, the physicochemical properties of the mine
tailings (MT) and sediments (SD1–6), the P1–6 paddy
soils and the P7–8 paddy soils were found to resemble
those of the Sp1, Sp2 and Sp3 soils, respectively (Table 3;
Fig. 1).

RESULTS

General physicochemical properties of the soils
The soil profile at the Sp1 site was highly disturbed as a
result of mining activity; the A horizon was mainly composed of colluvial materials from the upper slope, whereas
the BC and C horizons contained abundant rock fragments of various sizes. The soil profiles from the Sp2 and
Sp3 sites resembled each other. However, the profile at
the Sp2 site appeared to be more affected by redox reactions owing to uncontrolled flooding caused by the collapse of the dike. The Ap horizon was thicker at the Sp2
site than at the Sp3 site. It should be noted that the soil
matrix color was greenish gray or bluish gray throughout
the profile at the Sp2 site, but was brownish gray at the
Sp3 site.
The physicochemical properties were determined for
soils collected from three soil profiles in the Sp1, Sp2 and
Sp3 sites (Table 2). The Sp1 soils had a clayey texture
throughout the profile, with a higher content of TC in the
A horizon. The soil pH values were neutral in reaction,
with the highest Ex-Mg content among the exchangeable
bases. Compared with the Sp1 soils, the Sp3 soils had
lower clay contents and a more acidic nature, with higher
Ex-Ca contents, which were the highest among the
exchangeable bases. The Sp2 soils showed intermediate
properties between the Sp1 and Sp3 soils in terms of pH
and exchangeable bases. However, the clay content of the

Vertical distribution of total heavy metals and
As in the soils
The vertical distributions of total Cr, Co, Ni, Cu, Pb and
As were determined in the soil profiles at the Sp1, Sp2 and
Sp3 sites. There were different tendencies in the distribution patterns between Cr, Co and Ni and Cu, Pb and As
(Fig. 2).
The levels of Cr, Co and Ni were considerably higher in

the Sp1 soils than in the Sp3 soils, the respective ranges
were 1,616–4,331 mg kg)1, 162–286 mg kg)1 and
4,226–4,700 mg kg)1. At the Sp1 site, the content of Cr
and Co was higher in the A horizon than in the BC and C
horizons, whereas the Ni content was constant with
depth. In contrast, at the Sp3 site, levels of Cr, Co and Ni
did not vary appreciably throughout the profile. The Sp2

Table 2 Physicochemical properties of the soils taken from three soil profiles
TC
Site

Horizon

Mine soil
Sp1 A
BC
C
Paddy soils
Sp2 Ap
B1gir
B2gir
B3gir
Sp3 Ap
B1g
B2gir
B3gir

†

)1

TN

Ex-K

Ex-Na

)1

Ex-Ca

Ex-Mg

CEC

)1

Depth (cm)

pH

EC (mS m )

(g kg )

(cmolc kg )

0–20 ⁄ 33
20 ⁄ 33–40 ⁄ 50

50–68+

6.82
7.10
7.20

4.50
2.35
2.25

42.0
18.6
11.9

3.75
2.29
1.60

0.16
0.03
0.03

0.07
0.06
0.07

0.75
0.50
0.38

8.77
6.77
7.55

0–23
23–50
50–73
73–100+
0–9 ⁄ 10
9 ⁄ 10–33
33–57
57–95+

6.60
6.89
6.93
7.21
5.36
5.89
5.84
5.79

15.8
4.05
3.28
3.80
2.12
0.62
0.71
1.07

20.8
17.9
10.9
1.95
20.0
4.56
1.59
1.07

1.67
1.54
0.80
0.37
2.02
0.66
0.34
0.32

0.17
0.17
0.20
0.19
0.14
0.05
0.06
0.08

0.16
0.14

0.14
0.18
0.08
0.09
0.07
0.06

2.06
1.74
1.14
1.17
3.37
2.70
2.29
2.06

7.18
6.91
6.50
6.63
0.71
1.23
1.61
1.56

Clay (%)

Texture

32.6

36.4
43.6

39
40
41

CL
SC
SC

51.2
38.4
27.7
27.7
12.2
11.1
9.59
8.89

53
54
57
57
30
37
28
23

C

C
C
C
CL
CL
SCL
SCL

†

pH, pH(H2O). CEC, cation exchange capacity; C, clay ; CL, clay loam; EC, electrical conductivity; Ex-, exchangeable; SC, sandy clay; SCL, sandy clay
loam ; TC, total carbon; TN, total nitrogen.

Ó 2010 Japanese Society of Soil Science and Plant Nutrition


Soil contamination by a chromite mine in Vietnam

349

Table 3 Physicochemical properties of the mine tailings, sediments and paddy soils
TC
Site
Mine tailings
MT
Sediment
SD1
SD2
SD3
SD4

SD5
SD6
Paddy soil
P1
P2
P3
P4
P5
P6
P7
P8

†

)1

TN

Ex-K

Ex-Na

)1

Ex-Ca

Ex-Mg

Depth (cm)

pH

EC (mS m )

0–25

8.02

7.27

10.0

0.66

0.07

0.08

0.48

7.86

0–25
0–25
0–25
0–25
0–25
0–25

7.52

7.37
7.74
8.25
6.98
6.12

5.33
4.13
5.52
5.43
69.3
22.8

10.6
5.82
10.7
4.92
16.7
27.1

0.61
0.32
0.58
0.30
0.37
2.03

0.06
0.04
0.05

0.07
0.15
0.09

0.08
0.10
0.06
0.09
0.13
0.13

0.45
1.04
0.43
0.54
2.50
1.93

0–5
0–5
0–5
20–25
0–5
20–25
0–5
20–25
0–5
20–25
0–5
20–25

0–5
20–25

6.68
6.82
7.32
7.65
6.52
7.73
5.18
5.26
5.30
5.35
5.94
6.99
5.91
6.89

11.8
34.5
12.9
7.87
15.1
16.4
6.22
5.00
3.55
3.18
3.95
2.99

4.26
4.05

15.4
23.5
15.5
10.4
29.8
14.8
21.1
21.5
18.1
12.4
21.3
5.28
15.9
4.75

0.90
1.12
0.93
0.84
2.10
1.10
1.53
1.75
1.49
1.00
1.76
0.47

1.25
0.44

0.22
0.23
0.26
0.20
0.26
0.24
0.15
0.22
0.14
0.09
0.14
0.12
0.12
0.09

0.12
0.13
0.12
0.10
0.14
0.14
0.12
0.15
0.10
0.08
0.11
0.10

0.12
0.11

12.3

0.70

0.08

20.6
14.8

1.35
1.17

18.6
5.02

1.51
0.46

Average values
Mine tailings and sediments
(n = 7)
0–25
7.43
17.1
Paddy soils
Taken near mining sites (P1–P6)
(n = 6)

0–5
6.30
14.0
(n = 4)
20–25
6.50
8.11
Taken 5–6 km from the mining site (P7–P8)
(n = 2)
0–5
5.93
4.11
(n = 2)
20–25
6.94
3.52

CEC

)1

(g kg )

(cmolc kg )

Clay (%)

Texture

49.6

38

CL

4.84
6.77
5.85
6.85
3.86
6.99

24.5
28.9
22.7
23.4
36.4
44.5

12
24
12
7
19
54

SL
SCL
SL
S

SL
C

1.81
3.35
2.33
1.83
6.03
3.80
4.67
4.29
3.62
3.10
4.61
5.57
4.42
4.64

5.85
6.86
6.85
4.85
4.86
3.86
2.77
1.79
1.75
0.94
0.86
1.84

1.74
1.78

34.5
54.8
50.0
38.3
46.9
45.2
15.6
15.7
12.4
9.05
11.5
12.4
10.1
9.10

54
46
54
53
53
56
29
38
21
19
33
28

21
22

C
SIC
C
C
C
C
CL
CL
L
L
SCL
CL
L
L

0.10

1.05

6.15

32.9

24

0.21
0.19

0.12
0.12

3.64
3.26

4.82
2.86

35.7
27.1

43
42

0.13
0.11

0.12
0.11

4.52
5.11

1.30
1.81

10.8
10.8

27
25

†

pH, pH(H2O). CEC, cation exchange capacity; C, clay ; CL, clay loam; EC, electrical conductivity; Ex-, exchangeable; SC, sandy clay; SCL, sandy clay
loam ; TC, total carbon; TN, total nitrogen.

soils had intermediate levels of these heavy metals
between the Sp1 and Sp3 soils, but they were higher in the
Ap horizon than in the lower horizons.
In contrast to Cr, Co and Ni, the levels of Cu, Pb and
As were lower in the Sp1 soils than in the Sp3 soils. In the
Sp1 soils, Cu and As did not vary with depth, whereas Pb
content was high in the A horizon and decreased slightly
with depth. At the Sp3 site, the Cu content was highest in
the Ap horizon and almost constant in the deeper horizons, whereas Pb content had a minimum value in the
B2gir horizon. The total As content at this site was lower
in the Ap horizon than in the B1g to B3gir horizons. In
the case of the Sp2 site, although the Pb and As contents
in the Ap horizon were equivalent or similar to those at
the Sp1 site, these values increased with depth, and
reached similar levels to those of the Sp3 site in the deeper
Ó 2010 Japanese Society of Soil Science and Plant Nutrition

horizons. Similar to Pb and As, Cu at the Sp2 site tended
to be lowest in the Ap horizon, but with relatively high
levels throughout the profile.

Horizontal distribution of total heavy metals and
As in the soils
Figures 3 and 4 record the levels of total Cr, Co and Ni
and Cu, Pb and As, respectively, in the P1–8 soils, tailings
(MT) and sediments (SD1–6) collected from the sites
shown in Fig. 1.
High contents of Cr, Co and Ni were found in the MT
tailings (10,428, 340 and 4,200 mg kg)1, respectively)
and in the SD1–6 sediments (2,674–6,700, 294–1,000
and 3,477–7,800 mg kg)1, respectively). The paddy field
sites could be classified into three groups based on the
content of these heavy metals, which decreased signifi-


350 C. N. Kien et al.

Figure 2 Vertical distribution of heavy metals (Cr, Co, Ni, Cu and Pb) and As in three soil profiles (Sp1–Sp3).

cantly with increasing distance from P1 (Fig. 3). The first
group contained the P1–4 sites and the levels of these
heavy metals were high compared with the levels in the
sediment from SD6. The second group contained the P7
and P8 sites, where the soils were low in Cr, Co and Ni
content, similar to the Sp3 site. The P5 and P6 sites were
in the last group, and the levels of heavy metals at these
sites were higher than those in the second group, but were
considerably lower than those in the first group.
In contrast, the contents of Cu, Pb and As (ranges of
4.12–17.5, 7.14–17.2 and 1.35–4.79 mg kg)1, respectively) in the MT tailings and SD1–6 sediments were
almost comparable to the levels found in the Sp1 soils. In

the P1–8 sites, the paddy soil ranges were 14.4–64.9 mg
kg)1 for Cu, 14.4–93.1 mg kg)1 for Pb and 3.03–11.3 mg
kg)1 for As in the surface layers, and 21.2–41.6 mg kg)1
for Cu, 19.2–77.0 mg kg)1 for Pb and 3.41–13.8 mg kg)1
for As in the subsurface layers. The contents of these elements increased significantly with increasing distance
from P1 as illustrated by the correlation coefficients in
Fig. 4.

Water-extractable heavy metals and As
As a whole, small amounts of heavy metals and As were
detected in water-extractable form in the sediment

samples (SD2 and SD6) and soil samples (Sp1–Sp3) compared with their total content (Table 4).
The levels of Cr, Co and Ni in water-extractable form
were relatively high in the sediments and in the Sp1 and
Sp2 soils. These metals in the Sp1 and Sp2 soils decreased
with depth, except for Co in the Sp1 soils. The Sp3 soils
showed a similar distribution pattern for Cr, Co and Ni,
but the maximum levels found in the Ap horizon were
lower than the minimums found in the B3gir horizon for
the Sp2 site. It is noteworthy that the Co content of the
Sp1 and Sp2 soils was higher than that of the sediments,
whereas the Ni content was highest in the Ap horizon at
Sp2 site.
The content of water-extractable Cu and Pb in the Sp1
and Sp2 soils was higher than that in the SD2, SD6 and
Sp3 soils. Within the profile, Cu content was high in the
BC and C horizons at the Sp1 site, and in the B1gir and
B2gir horizons at the Sp2 site, whereas Pb content was
high in the A horizon at the Sp1 site and in the B1gir

and B2gir horizons at the Sp2 site. At the Sp3 site, Cu
and Pb tended to be high in the Ap and B1g horizons.
Water-extractable As was not detected in the SD2 and
SD6 sediments or in the Sp1 soils, but soils in the upper
two horizons at the Sp2 site showed negligible content of
As.
Ó 2010 Japanese Society of Soil Science and Plant Nutrition


Soil contamination by a chromite mine in Vietnam

351

Figure 3 (A–C) Horizontal distribution of Cr, Co and Ni in the mine tailings, sediments and paddy soils. , surface layer; , subsurface
layer. (a–c) Relationships between Cr, Co and Ni in the surface layer of the paddy soils and the distance from the P1 site.

Ó 2010 Japanese Society of Soil Science and Plant Nutrition


352 C. N. Kien et al.

Figure 4 (A–C) Horizontal distribution of Cu, Pb and As in the mine tailings, sediments and paddy soils. , surface layer; , subsurface
layer. (a–c) Relationships between Cr, Co and Ni in the surface layer of the paddy soils and the distance from the P1 site.
Ó 2010 Japanese Society of Soil Science and Plant Nutrition


Soil contamination by a chromite mine in Vietnam

353

Table 4 Water-extractable contents of heavy metals and As in the sediments and soils taken from three soil profiles
Cr
Site

Horizon

Sediments
SD2
SD6
Mine soil
Sp1

Ap
B1gir
B2gir
B3gir
Ap
B1g
B2gir
B3gir

Sp3

Ni

Cu

Pb

As

(mg kg)1)

Depth (cm)

A
BC
C

Paddy soils
Sp2

Co

0–25
0–25

11.4
12.1

0.26
0.38

17.7
24.4

0.04
0.11

0.04

0.28

–
–

0–20 ⁄ 33
20 ⁄ 33–40 ⁄ 50
50–68+

13.2
8.81
3.55

0.96
1.65
3.06

21.1
14.6
10.4

0.21
0.53
0.81

0.21
0.08
0.04

–

–
–

0–23
23–50
50–73
73–100+
0–9 ⁄ 10
9 ⁄ 10–33
33–57
57–95+

12.7
8.63
3.02
0.67
0.48
0.13
0.06
0.06

1.16
1.00
0.72
0.07
0.02
0.01
–
–

32.3
10.5
2.65
1.06
0.32
0.13
0.06
0.06

0.22
0.55
0.58
0.02
0.16
0.13
0.03
0.01

0.32
1.08
1.01
–
0.12
0.20
0.03
0.02

0.05
0.06
–

–
0.15
0.17
0.10
–

Concentrations are calculated as the element content in the dry mass of the sample. –, not detected.

Hexavalent chromium
Taking into consideration that the toxicity of Cr(VI) is
much higher than that of Cr(III) (Adriano 2001), the
amount of total Cr(VI) was determined for the soil samples from the Sp1, Sp2 and Sp3 sites and for the sediment
samples SD2 and SD6 (Table 5). Table 5 also gives the
results of field measurements of the pH and Eh values.
Table 5 pH, redox potential and hexavalent chromium content
in the sediments and soils taken from three soil profiles

Site

Horizon

Sediments
SD2
SD6
Mine soil
Sp1

A
BC
C

Paddy soils
Sp2
Ap
B1gir
B2gir
B3gir
Sp3
Ap
B1g
B2gir
B3gir

pH†

Eh
(mV)

Cr(VI)
(mg kg)1)

0–25
0–25

7.97
6.76

43
)208

2.47
10.1

0–20 ⁄ 33
20 ⁄ 33–40 ⁄ 50
50–68+

6.61
7.37
7.59

242
236
212

36.0
22.5
9.54

0–23
23–50
50–73
73–100+
0–9 ⁄ 10
9 ⁄ 10–33
33–57
57–95+

7.40
7.27

7.38
7.60
5.52
5.80
6.20
6.13

24
124
85
141
276
287
260
252

9.79
5.34
1.71
0.50
1.25
0.74
0.30
0.11

Depth
(cm)

†
pH, pH (H2O). Concentrations are calculated as the element content in

the dry mass of the sample. Cr(VI), hexavalent chromium; Eh, redox
potential.

Ó 2010 Japanese Society of Soil Science and Plant Nutrition

No clear relationships were found between the content
of Cr(VI) and pH or Eh values. The Cr(VI) content was
higher at the Sp1 and Sp2 sites than at the Sp3 site. The
vertical distribution of Cr(VI) in the soil profile resembled
that of water-extractable Cr. However, the Sp1 soils had
much higher levels of Cr(VI) than water-extractable Cr,
although the levels of these Cr fractions were similar to
each other in the Sp2 soils.

Concentration of heavy metals and As in the
water samples
In the water samples, as distance from the mining area
increased, higher concentrations of certain elements were
found; for example, moving from the MT to the P2 site
for Cr and moving from the SD1 to the P4 site for Ni
(Table 6). The concentration of Ni, in particular, often
exceeded the regulation limit of 100 lg L)1 set by the
‘‘Vietnamese standard limitation for surface water’’
(TCVN 5942-1995 1995). In contrast, Co, Cu, Pb and As
were not detected.

DISCUSSION
Influence of mining activity on adjacent paddy
fields
Based on the results related to the vertical and horizontal

distribution patterns of total heavy metals and As, as well
as the general physicochemical soil properties and the soil
profile descriptions, mining activity caused Cr, Co and
Ni contamination of soils in adjacent paddy fields by
sediment inflow at the time of the dike collapse, which


354 C. N. Kien et al.

Table 6 pH, electrical conductivity and concentration of heavy metals and As in the water samples
Cr

Co

Ni

Cu

Pb

As

Site

Sample type

pH

EC (mS m)1)

(lg L)1)

MT
SD1
SD2†
SD3
SD4
SD5
SD6†
P1
P2
P3†
P4
P5
P6
P7
P8
Vietnamese standard
limitation for surface
water (TCVN 5942-1995 1995)

Stream water
Stream water

8.20
7.90

2.89
2.75

32.0
15.0

–
–

27.0
129

–
–

–
–

–
–

Stream water
Stream water
Lake water

7.70
7.80
7.90

2.15
2.83
2.39

17.0
25.0
41.0

–
3.00
–

135
123
20.0

–
–
–

–
–
–

–
–
–

Standing water
Standing water

7.50
7.30

1.24
0.92

8.00
75.0

–
–

87.0
174

–
–

–
–

–
–

Standing water
Standing water
Standing water
Standing water
Standing water

7.20
7.00
6.50

7.00
6.90

0.99
0.64
0.66
0.86
1.29

9.00
10.0
9.00
3.00
4.00
50 Cr(VI)

–
–
4
–
3
100

–
–
–
–
–
50

–
–
–
–
–
10

–
–
–
–
–
N.A.

132
36.0
35.0
8.00
15.0
100

†

Water samples were not taken at the SD2, SD6 and P3 sites. EC, electrical conductivity; –, not detected; N.A., not available.

happened as a result of flooding. This contamination
faded away with distance from the dike, but extended for
approximately 2.0 km (to the P5 and P6 sites; Fig. 3).
Levels of these heavy metals were considerably higher
throughout the soil profile at the Sp2 site, particularly in

the Ap horizon, than the levels at the Sp3 site (Fig. 2),
indicating that the collapse of the dike and the resulting
inflow of sediments had occurred repeatedly, and that
such sediments had been mixed with the original soils
through agricultural works such as puddling. Taking into
consideration the significant concentrations of Cr and Ni
in the water samples, water inflow from the breaking dike
and irrigation water, if provided from the mining area,
might also contribute to Cr and Ni contamination of the
soils (Table 6).
In contrast to the Cr, Co and Ni contamination, which
resulted from mining activity, levels of Cu, Pb and As
were low in the soil, sediment and tailing samples in the
mining area. Levels of these elements in the lowland area
appeared to depend on the constituents of the parent
materials (Adriano 2001; Vaselli et al. 1997) because the
total contents of these elements were highest at the Sp3
site, and because the B1gir and B2gir horizons at the Sp2
site showed comparable levels of these elements to the levels recorded at the Sp3 site (Fig. 2). At the Sp3 site, the
levels of water-extractable Cu and Pb were high in the
upper two horizons (Ap and B1g), which could be
ascribed to organic matter accumulation derived from
plant residue and manure (Adriano 2001). However, at
the Sp2 site, water-extractable Cu and Pb from the soils
showed a peak in content in the B1gir and B2gir horizons,

and these levels were higher than at the Sp3 site, suggesting a different source for these heavy metals. Although the
content of water-extractable Cu and Pb in sediments at
the SD2 and SD6 sites was low, one possible explanation
applicable for Cu, and probably for Pb, is the heterogeneity of these constituents in the mine ores, tailings and sediments within the mining area. This was demonstrated by

the different vertical distribution patterns of the waterextractable fractions of heavy metals at the Sp1 site, that
is, the decreasing trends with depth for Cr, Ni and Pb and
the increasing trends with depth for Co and Cu.

Potential health hazards as a result of
contamination
Paddy fields located within 2.0 km of the dike were contaminated with Cr, Co and Ni, although the contamination faded with distance. The total content of Cr, Co
and Ni at the P1–P6 sites greatly exceeded the normal
levels in uncontaminated soils reported by Bowen
(1979), that is, 70 mg kg)1 for Cr, 8 mg kg)1 for Co
and 50 mg kg)1 for Ni. We have previously surveyed
paddy soils near the tin and tungsten mines of Daitu district, Vietnam, and established that the total Cr and Ni
ranges were 20.8–27.3 and 5.67–7.48 mg kg)1, respectively (Kien et al. 2009). In contrast, Phuong et al.
(2009) examined metal pollution of paddy soils in the
Red River Delta caused by industrial activities and natural sources and found that the total Cr content was 23–
136 mg kg)1 for the surface soils. The total contents of
Cr and Ni at the P1–P4 sites of the present study were
very much higher than the above-mentioned values. The
Ó 2010 Japanese Society of Soil Science and Plant Nutrition


Soil contamination by a chromite mine in Vietnam

severe contamination found in the present study can be
clearly ascribed to the dike collapse, which allowed
direct inflow into the paddy fields of sediments that had
a high content of heavy metals.
Water-extractable fractions of Cr, Ni and Co from the
soils were higher at the Sp1 and Sp2 sites than at the Sp3
site (Table 4). The Ap horizon at the Sp2 site in particular

had the highest Ni value among all soil and sediment samples. According to a review on the behavior of Ni and Cr
in soils (McGrath and Smith 1990), Ni is more soluble in
water and more easily mobile through soils than Cr.
Although accounting for only a small percentage of the
total content, the amounts of Cr, Ni and Co in the Ap
horizon were 26, 58 and 100-fold higher at the Sp2 site
than at the Sp3 site. Because water-extractable metals represent the fraction weakly bound to the soil matrix and,
therefore, are readily available to plants (Abollino et al.
2002), it is possible that the rice plants absorb and accumulate this fraction, leading to the occurrence of a serious
health hazard.
A small portion of the total Cr was extracted as the
highly toxic Cr(VI), probably because a major form of Cr
in the mining area is Cr2O3 minerals, and a small amount
of Cr(VI) can be released into soils during the weathering
process (Table 5). However, despite the low Eh value, the
level of Cr(VI) in the Ap horizon at the Sp2 site was relatively high, within phytotoxic levels. Turner and Rust
(1971) reported that a Cr(VI) level of approximately
0.5 mg L)1 in solution and 5 mg kg)1 in soils could be
toxic to plants. Adema and Henzen (1989) evaluated the
effect of Cr(VI) on the growth of lettuce (Lactuca sativa),
tomato (Lycopersicon esculentum) and oats (Avena sativa) grown in humic sandy and loamy soils and reported
that the growth of the lettuce, tomato and oats was inhibited, resulting in 50% yield reductions, at levels of 1.8,
6.8 and 7.4 mg kg)1 Cr(VI) in the soils, respectively. The
total root weight and root length of wheat was affected by
20 mg kg)1 Cr(VI) in the soil as K2Cr2O7 (Chen et al.
2001).
In addition to sediment inflow, water flowing from the
mining area and used for irrigation and drinking can also
be a health hazard owing to the high concentrations of Cr
and Ni. Furthermore, in this area polluted water would

affect domestic animals, such as water buffalo, cattle and
ducks, which are bred in and around the paddy fields for
farm work or for food.
In contrast, the mining activities in the surveyed area in
the present study would not cause severe contamination
of Cu, Pb and As because the levels of these elements were
below their typical levels in uncontaminated soil (30 mg
kg)1 for Cu, 35 mg Pb kg)1 and 6 mg kg)1 for As), as
reported by Bowen (1979). However, these levels were
comparable to the maximum allowable limit for agricultural soils set by the Vietnamese government (TCVN
Ó 2010 Japanese Society of Soil Science and Plant Nutrition

355

7209-2002 2002; 50 mg kg)1 for Cu, 70 mg Pb kg)1 and
12 mg kg)1for As). In addition, the data support the suspicion that some portion of the Cu and Pb might originate
from the sediments of the mine (Table 4). Therefore, the
amounts and chemical forms of Cu, Pb and As, as well as
Cr, Co and Ni, should be monitored regularly.

Conclusions
The present study revealed that mining activity at the Co
Dinh chromite mine caused severe contamination of lowland paddy fields with Cr, Ni and Co, which might cause
serious health hazards through agricultural products,
including livestock. It is clear that the principal source of
the pollution was sediment inflow as a result of the collapse of the dike, which was poorly constructed by heaping up soil. Because the affected area was limited, at most
2 km from the dike, appropriate countermeasures could
be applied, probably without difficulty if official support
was available. For example, although a solid dike built
with concrete was constructed to prevent further inflow of

sediments, the polluted area should be enclosed until
appropriate decontamination measures have been applied.
However, water from the mining area was also polluted
with Cr and Ni. This might be another problem because
the lake water flows into a river (Fig. 1). Although the
present study provided fundamental data related to water
quality, further monitoring is required, including monitoring fluctuations in water quality with seasonal changes.
Furthermore, monitoring of changes in the chemical forms
of the contaminants in the agricultural soils, as well as in
their concentrations in the crops (e.g. rice plants), should
be carried out regularly.

ACKNOWLEDGMENTS
This research was financially supported by a Grant-in-Aid
for Scientific Research to K. Iwasaki (grant no.
18380195) from the Japan Society for the Promotion of
Science. The authors sincerely thank the officers at the
sampling sites and our colleagues at the Hanoi University
of Science, Vietnam, for their valuable help and support
during sample collection.

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Ó 2010 Japanese Society of Soil Science and Plant Nutrition