Environ Monit Assess (2014) 186:1499–1513 DOI 10.1007/s10661-013-3469-8
State of rare earth elements in different environmental components in mining areas of China Tao Liang & Kexin Li & Lingqing Wang
Received: 16 May 2013 / Accepted: 1 October 2013 / Published online: 18 October 2013 # Springer Science+Business Media Dordrecht 2013
Abstract China has relatively abundant rare earth elements (REEs) reserves and will continue to be one of the major producers of REEs for the world market in the foreseeable future. However, due to the large scale of mining and refining activities, large amounts of REEs have been released to the surrounding environment and caused harmful effects on local residents. This paper summarizes the data about the contents and translocation of REEs in soils, waters, atmosphere, and plants in REE mining areas of China and discusses the characteristics of their forms, distribution, fractionation, and influencing factors. Obviously high concentrations of REEs with active and bioavailable forms are observed in all environmental media. The mobility and bioavailability of REEs are enhanced. The distribution patterns of REEs in soils and water bodies are all in line with their parent rocks. Significant fractionation phenomenon among individual members of REEs was found in soil–plant systems. However, limited knowledge was available for REEs in atmosphere. More studies focusing on the behavior of REEs in ambient air of REE mining areas in China are highly suggested. In addition, systematic study on the translocation and circulation of REEs in various media in REEs mining areas and their health risk assessment should be carried out. Standard analytical methods of REEs in environments need to be
T. Liang (*) : K. Li : L. Wang Institute of Geographical Sciences and Natural Resources Research, CAS, Beijing 100101, China e-mail: [emailprotected]
established, and more specific guideline values of REEs in foods should also be developed. Keywords Rare earth element . Chinese REE mining areas . Environmental behavior
Introduction It is widely recognized that mining activities have changed the original environmental condition and produced vast environmental problems, including ecological destruction, environmental pollution, soil erosion, and geological disasters (Aguilar et al. 2004; Klukanová et al. 1999; Liu et al. 2006; Luís et al. 2011; Salomons 1995). More toxic elements may be released into the environment and cause adverse impacts because of the disturbed environmental condition. During the last decades, a large number of studies have focused on the geochemistry of heavy metals and metalloids whose toxic effects have been well-understood in mining areas, such as As and Hg (Natarajan et al. 2006; Teng et al. 2004), whereas fewer researches have paid attention to the behavior of rare earth elements (REEs). However, with the development of mining activities and the increasing use of REEs as a new material in recent years, more studies have been paid on REEs .Several studies have indicated the harmful effects and health hazards of REEs to human beings, and it has already been proven that long-term exposure of REE dust may cause pneumoconiosis in humans (Hirano et al. 1996). In this paper, the geochemistry including several aspects, such as
1500
concentration, distribution pattern, and controlling factors of REEs in soils, waters, and plants in REE mining areas of China will be summarized in order to provide a better understanding of the environmental behavior of REEs.
Outline of Chinese REE resources and industry REEs lie at the bottom of Mendeleev’s Periodic Table, including 15 lanthanides, from lanthanum (La) to lutetium (Lu), scandium (Sc), and yttrium (Y). The 17 elements form a coherent group of elements with similar chemical properties. Commonly, REEs can be divided into three groups by their atomic number and masses—the light rare earth elements (LREE), the middle rare earth elements (MREE), and the heavy rare earth elements (HREE). La, Ce, and Pr are considered to be LREEs; Nd, Sm, Eu, and Gd are classified to MREEs, and those from Tb to Lu are in the group of HREEs (EPA 2012) China has relatively abundant REE resources, and its RE reserve accounts for 23 % of world’s proven reserve (18.59Mt, in 2009). The distribution of REE resources in China presents a “light north, heavy south” pattern (Anonymous 2012). LREE mines mainly exist in Baotou of the Inner Mongolia Autonomous Region, as well as in Liangshan of Sichuan province. MREE and HREE mines are located in Ganzhou of Jiangxi province. Despite of the plenty reserves, China’s REE resources also show a rather high diversity in types of rare earth minerals which include bastnaesite, monazite, ion-absorption minerals, xenotime, fergusonite, and others, with a relatively complete range of REEs (Cheng et al. 2010). It is worth mentioning that the ion-absorption MREE and HREE ores are China’s unique REE resources with an important position in the world (Li et al. 2010). Since China’s reform and open-door policy in 1997, China’s rare earth industries have developed rapidly. It is reported that with a complete technological system in mining, dressing, smelting, and separating of rare earth ores, China can produce over 400 varieties of rare earth products in more than 1,000 specifications. In 2011, China has produced over 90 % of the world’s rare earth supply with the only 23 % of the world total reserves and has become to a leading producer of REEs (Anonymous 2012). However, such a large scale of mining and production
Environ Monit Assess (2014) 186:1499–1513
activities of REEs in China has inevitably changed the environment.
High level of REEs in China’s rare earth mining areas REEs getting “rare” in their name is because of their scattered distribution and difficulty in the refining process. The term “rare” in name of REEs do not imply that they are uncommon in nature. On the contrast, REEs are relatively abundant in the earth. The total contents of REEs exceed 200 ppm in the average crust. Some REEs are even more common than copper or lead in the crust (Castor and James 2006; Chen 2011). Table 1 shows the abundances of REEs in the upper crust. In China, even though the background value is relatively high, REEs existing in rocks under natural conditions normally have little effects on the environment and human beings. However, in recent years, with the increased exploitation of REEs in China, mining activities have led to intensive accumulation of REEs and enhanced the level of REEs in the environment high enough to cause harmful effects on human beings (Feng et al. 2005; Husain et al. 1980). Many Chinese researches have carried out on the extent of the accumulation of exogenous REEs in China’s REE mining areas. Surface soil REEs are naturally omnipresent in soils which are mainly originated from the weathering of local parent rocks. Under natural conditions, the concentration of REEs in soils is mainly decided by their parent rocks (Hu et al. 2006). But in REE mining areas where natural conditions have been disturbed intensively, soils always have much higher concentrations compared with their parent rocks. Concentrations of REEs in several REE mining area in China are summarized in Table 2. As shown in the Table, the total contents of REEs in soils close to mining areas are all much higher than the average values in China and in the world. Soils in Bayan Obo REE Deposit show the highest concentration of REEs which reached 27,549.58 μg/g (Li et al. 2008). This concentration is almost 160 times higher than the average value in China. In southern REE mining areas of China, the contents of REEs are also high which range from 396 to 2,314 μg/g, and even the lowest concentration of
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Table 1 The upper crust abundances of REEs (Taylor and Mclennan 1985) Element
La
Ce
Pr
Nd
Pm
Sm
Eu
Gd
Tb
Dy
Ho
Er
Th
Yb
Lu
Abundance (ppm)
30
64
7.1
26
na
4.5
0.88
3.8
0.64
3.5
0.80
2.3
0.33
2.2
0.32
REEs in the soils is twice higher than the Chinese average concentration. Water body Fresh water normally contains very low concentrations of REEs. The concentrations of REEs in water only range between n×10−3 ∼n×10−1 μg/L (Elderfield et al. 1990; Ji et al. 2004; Sholkvitz 1993, 1995; Sultan et al. 2009). Due to the mining activities, large amounts of REEs are released into natural water bodies in two ways: the direct discharge of industrial wastewater which contains high levels of REEs and the leaching process in soils which are enriched in REEs. As a result, concentrations of REEs in water bodies have been significantly enhanced. Yellow river, the second biggest river in China, is the best example (He et al. 2004a, b, c). Baotou city where Bayan Obo REE
Deposit is located is famous for its REE industry. The Baotou section of the Yellow River received most of the wastewater from REE industries in Baotou. Table 3 shows the concentration of REEs in the mainstream and two branches (Kundulun River and Sidaosha River) of Baotou section of the Yellow River. It can be seen from Table 3 that the concentrations of REEs in surface water, as well as the suspended particles and surface sediment in Baotou section of the Yellow River, are all significantly higher than the average value of rivers in North China (Zhou et al. 2012). The mainstream had the lowest total REE concentration in comparison with the two branches. Sidaosha River had the highest level of REEs. The total concentration of REEs in suspended particles and surface sediments of Sidaosha River are 31,524 and 30,461 μg/L, respectively. Both of them are almost 200 times higher than the average value of rivers in North China.
Table 2 Concentrations of REEs in soil samples in REE mining areas of China (micrograms per gram) Place
Jiangxi REE mining areas
Baotou REE mining areas
China average
World average
pH
–
5.26
3.69
4.13
4.30
7.02
7.02
–
–
La
208.50
62.3
110
492
30
6,905.24
55.03
37.4
50
Ce
351.63
85.2
44.2
630
49
12,169.89
132.60
64.7
7
Pr
46.90
–
31.3
166
11
1,614.28
12.02
6.67
35
Nd
159.75
97
154
411
39
5,726.56
42.93
25.1
4.5
Sm
24.98
39.0
49.1
127
21
492.35
6.70
4.94
1
Eu
2.36
1.50
3.99
12
1
96.90
1.25
0.98
4
Gd
21.68
–
42.3
106
21
329.15
5.43
4.38
0.7
Tb
2.54
10.2
6.64
13
4
38.06
0.76
0.58
5
Dy
11.09
–
37.4
42
21
54.63
4.37
3.93
0.6
Ho
1.97
–
7.28
16
7
–
0.91
0.830
2
Er
5.47
–
19.7
33
17
–
2.54
2.42
0.6
Tm
2.87
–
2.45
7
3
–
0.38
0.35
3
Yb
4.58
36.7
15.9
27
20
–
2.53
2.32
0.4
Lu
0.68
5.60
1.91
6
3
–
0.40
0.35
40
Y
51.23
–
–
–
–
–
–
–
–
∑REE
869.73
542
688.2
2314
396
27,549.58
279.84
176.8
193.8
Reference
Zhu et al. (2002)
Gao et al. (2001)
Yu et al. (2009)
Wei et al. (2002)
Wei et al. (2002)
Li et al. (2001)
Zhang et al. (2012)
Wei et al. (1991)
Wang et al. (1989)
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Table 3 Concentrations of REEs in Huanghe River (Baotou section) La
Ce
Pr
Nd
Sm
Eu
Gd
Tb
Dy
Ho
Er
Tm
Yb
Lu
∑REE
Dissolved REE(μg/L)a Yellow river 0.10 (main stream) Kundulun River 140
0.22
0.034 0.095 0.054 0.018 0.028 0.006 0.08
0.014 0.03
0.007 0.032 0.006 0.80
152
16.08 52.01 6.91
Sidaosha river
1,149
294
Southern river 0.065 averageb Suspended matter (μg/g)c
0.135
0.014 0.065 0.014 0.016 0.011 0.002 0.011 0.003 0.014 0.002 0.013 0.002 0.367
Yellow river 33.16 (main stream) Kundulun River 86.16
68.26
8.47
28.70 5.79
1.31
6.29
0.82
3.39
0.77
2.25
0.25
1.76
0.25
161
139.3
15.8
51.98 6.61
1.20
4.31
0.97
1.87
0.68
1.36
0.22
1.01
0.22
312
Sidaosha river
7,746
14,472 1,162 6,901 478
98.59 423
63.38 88.03 12.32 33.45 5.63
35.21 5.28
31,524
Southern river averageb Sediment (μg/g)d
40.07
71.50
8.41
30.80 5.36
1.03
4.72
0.71
3.31
0.67
1.78
0.28
1.68
0.88
171.26
Yellow river 27.47 (main stream) Kundulun River 50.77
55.94
6.74
24.20 4.48
0.9
4.15
0.50
2.50
0.50
1.50
0.14
1.30
0.16
130
116.6
a
988
1.52
1,193 62.35 5.59
0.88
3.80
0.18
2.34
0.24
1.48
0.21
406
67.33 5.58
7.22
9.08
1.35
9.79
0.36
1.92
0.22
3,826
13.30 47.00 7.04
1.26
6.79
1.01
3.52
0.73
2.01
0.24
1.18
0.22
251
Sidaosha river
7,823.0 14,074 1,403 5,919 497
99.3
423
43.5
82.4
8.97
38.6
6.06
38.3
5.98
30,461
Southern river averageb
51.00
1.29
6.81
1.10
5.90
1.19
3.37
0.57
3.51
0.55
229.70
94.00
11.50 41.00 7.60
He et al. (2004a)
b
Zhou et al. (2012)
c
He et al. (2004b)
d
He et al. (2004c)
Additionally, it is well known that the concentrations of dissolved REEs in water under natural environment are limited which are often ignored (Ding et al. 2005a; Wang et al. 2000b). However, the total concentrations of dissolved REEs in rivers near REE mining areas in China are about three orders of magnitude higher than those in other natural rivers which are not affected by mining activities worldwide (Table 4). Leaching process in soils is another important pathway for exogenous REEs entering water bodies which could cause contamination in groundwater, especially in southern REE mining areas of China. The highest reported REE concentration in well water in Chinese mining areas is 130 μg/L which occurred in one of the exploratory mine of Shandong REE mining areas (Lu et al. 1995). Except for that extreme value, other reported REE concentrations in well water were not too high. Several published results of REE concentrations in groundwater are listed in Table 5. Even though the total concentrations in groundwater are not too high when compared with Baotou section of the Yellow River mentioned earlier, it
could lead to even worse effect because they are sources of drinking water for local people. Atmosphere Along with intensive worldwide use of REEs, an increasing amount of REEs has been emitted into atmosphere. Unfortunately, unlike in soils and water bodies where the contents of REEs have been widely studied, the levels of REEs in the air have not received enough attention. In general, existing studies have investigated three aspects: (1) the long haul transport of REEs from continents to oceans (Arimoto et al. 1989; Sholkovitz et al. 1993; Sholkovitz 1994a; Yang et al. 1993); (2) airborne particulate matter in urban areas and levels of REEs in the air (Chen et al. 1985; Hong et al. 2010; Suzuki et al. 2011; Wang 2003); and (3) air pollution caused by REEs in forms of ambient particles produced during the cracking and combustion of oil in the vicinity of refinery factories because of the fact that many petroleum-cracking catalyst and products are highly enriched in the LREEs (La, Ce,
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Table 4 Concentration of dissolved REEs in rivers μg/L
River’s name
La
Ce
Mining areas
Ganjiang Rivera
2.412
0.022
0.060
0.051
0.013
0.007
0.093
0.004
–
Huanghe River (Baotou section)b
0.10
0.22
0.095
0.054
0.018
0.006
0.032
0.006
0.800
Zhujiang River (Guangzhou section)c
0.035
0.075
0.050
0.037
0.015
0.010
0.023
0.005
0.246
Changjiang Riverd
0.050
0.110
0.070
0.013
0.004
0.008
0.002
0.261
China’s average e
0.060
0.120
0.060
0.010
0.005
0.003
0.008
0.002
0.268
Amazon Riverf
0.074
0.212
0.127
0.035
0.008
–
0.015
–
–
Indus River (ng/L)g
2.91
2.41
3.20
0.71
0.22
–
0.94
0.17
–
Natural areas
a
Nd
Sm
Eu
Tb
Yb
Lu
∑DREE
Meng et al. (2007)
b
He et al. (2004a)
c
Wang et al. (1998)
d
Zhang et al. (1992)
e
Wang et al. (1995)
f
Hu and Jin (2000)
g
Meng et al. 2008
and Nd) (Brown et al. 1991; Kitto et al. 1992; Olmez et al. 1985). There were only two studies about the REE mining areas, and both of them were carried out in Jiangxi province, China. Published results of REE concentrations in the air are listed in Table 6. Additionally, considering the adverse effects of atmospheric REEs on human beings, Zhu et al. (1996) investigated the concentrations of inhalable REEs in the air in an ion-adsorbed REE mining area in Jiangxi province. The results revealed a much higher level of REEs in inhalable particles in the mining area of 612.5 ng/m3 in comparison with REEs concentration in total suspended particulates in the same area of 359.1 ng/m3, indicating negative health effects on local residents. Plant tissues REEs are not essential elements to plants. However, plants can absorb REEs from soils through roots and can also uptake REEs via leaves exposed to atmospheric particles containing REEs mainly from REEs mining
areas (Ichihashi et al. 1992; Tyler 2004; Volokh et al. 1990). The concentrations of REEs in common plants are generally low under natural conditions, normally around 10−3–10−1 μg/g (dry mass) (Wang et al. 1997). But in the REE mining areas of China, both soils and water bodies have high levels of REEs as a result of mining activities as previously mentioned. Therefore, much attention has been paid on the investigation of the REE levels in plants grown in REE mining areas. Table 7 lists the REE concentrations in several previously investigated plants in REE mining areas of China. The results show that the level of REEs varied in different species of plants. Green stuff and sweet potato had high REE concentrations which are 10–20 times higher than the maximum limited levels of contaminants in foods issued in National Standards GB 2762-2005( Chinese Ministry of Health 2005). The total concentrations of REEs in all the other plants in the REE mining areas were lower than the limited level. Researchers have also carried out studies on REE hyperaccumulating plants in order to find a feasible
Table 5 Concentrations of REEs in groundwater in REE mining areas of China (micrograms per liter) Location
Deposit type
REEs mining areas
Jiangxi Jiangxi
Light REE Heavy REE
15.2±5.6 8.4±5.9
Control areas 7.0±3.7 5.6±2.2
Shandong
Light REE
5.1±2.6
3.2±0.8
Jiangxi
Light REE
9.18±5.82
0.38±0.28
Method
Reference
Spectrophotography
Lu et al. (1995)
ICP-MS
Zhu et al. (2002)
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Environ Monit Assess (2014) 186:1499–1513
Table 6 Concentrations of REEs in TSP ng/m3
Non-mining areas
Sampling location
Delft, Germany
Osaka, Japan
La
0.504
6.48
17.03
20.8
1,210
120.1
Ce
0.592
10.62
38.7
29.5
115
43.8
REE mining areas Beijing, China
Jiangxi, China
Jiangxi, China
Jiangxi, China
Pr
0.109
1.368
4.09
4.5
–
26.5
Nd
0.345
3.24
13.50
17.0
830
91.0
Sm
0.0635
0.468
2.82
2.7
119
17.1
Eu
0.0182
0.0378
0.59
0.3
125
1.9
Gd
0.0698
0.378
2.57
2.3
–
14.5
Tb
0.0084
0.0378
0.33
0.3
13.30
1.8
Dy
0.0460
0.342
1.63
1.4
–
8.3
Ho
0.0086
0.0594
0.29
0.2
–
1.2
Er
0.0234
0.1296
0.82
0.6
–
3.2
Tm
0.0030
0.01368
0.11
0.0
–
0.3
Yb
0.0235
0.135
0.65
0.5
13.5
2.0
Lu
0.0027
0.027
0.09
0.0
24.2
0.3 27.3
Y
0.294
–
3.28
6.5
–
∑REE
4.35
23.39
94.65
86.6
2,450
359.1
Method
ICP-MS
SSMS
ICP-MS
ICP-MS
–
ICP-MS
Reference
Wang et al. (2000a)
Sugimae (1980)
Wang et al. (2001a)
Zhu et al. (2003)
Peng and Wang (1995)
Zhu et al. (2003)
way to remediate REE-contaminated soils. REE hyperaccumulators can accumulate much higher contents of REEs, especially in their aboveground parts. The REE contents in the aboveground parts of plants are usually used as an indicator in the identification of REE hyperaccumulators which need to be higher than 1,000 μg/g (dry mass) (Garbisu et al. 2001). According to the criterion, four species have been widely recognized as REE hyperaccumulators (Table 8). In those four REE hyperaccumulators, Dicranopteris dichotoma is the most general and important REE Table 7 Concentrations of REEs in plants
Plant
Green stuff average Sweet potato Rice
REE mining areas
accumulating plant in REE mining areas in China which has been considered as an ideal material on the study of REE behavior in plants for many years (Wei et al. 2006). In China, Li et al. (1992) first found that D. dichotoma could accumulate REEs in southern part of Jiangxi province and reported the contents of REEs in different parts of D. dichotoma. Since then, Chinese scholars have been carrying out lots of researches on contents, forms, and distribution of REEs inside D. dichotoma. Several published results focusing on D. dichotoma grown in Chinese REE mining areas are summarized
Maximum limited level in food issue
Method
Reference
7.54
≤0.7
ICP-MS
Zhu et al. (2002)
13.59
≤0.5
Spectrophotography
Zhang et al. (1999)
1.55
≤2.0
Spectrophotography
Zhang et al. (1999)
Peanuts
0.36
≤0.5
Spectrophotography
Zhu et al. (1997)
Navel Orange
0.254
≤0.7
ICP-MS
Yu et al. (2009)
Tea leaf
1.59
≤2.0
ICP-MS
Zhang et al. (1997)
Environ Monit Assess (2014) 186:1499–1513 Table 8 Concentrations of REEs in leaves of four typical REE hyperaccumulator plants (micrograms per gram)
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Latin name of plant
Family of plant
Max. contents of REEs in leaves
Reference
Mockernut hickory
Juglandaceae
1,350
Thomas (1975)
Cary tomentosa
Juglandaceae
2,296
Robinson (1943)
Blechnum orientale
Blechnaceae
1,022
Xiao et al. (2003)
Dicraneopteris dicthotoma
Gleicheniaceae
3,358
Xiao et al. (2003)
in Table 9. Obviously, concentrations of REEs in leaves are in a high level. In summary, based on existing research results, we briefly present the biogeochemical cycle of REEs in Chinese REE mining areas (see Fig. 1). The range of total REE contents in different environmental media varied considerably. The levels of REEs in REE mining areas of China are obviously higher than those in non-mining areas.
However, with the growing amount of experimental results which indicated harmful effects of REEs on human beings, it is important to focus on the toxic effects of REEs, especially in REE mining areas. Investigations have been carried out on forms, distribution, and fractionation of REEs in soils, waters, atmosphere, and plants in different REE mining areas in China. Although all those published results are important for a better understanding of translocation of REEs in special areas, there is still lack of integrity and comparability.
Translocation of REEs in REE mining areas of China and influencing factors
Forms of REEs in Chinese mining areas
The translocation of REEs in various environmental compartments has gained considerable attention world widely over the last decade (Hu et al. 2006; Milliman et al. 1983; Sholkovitz 1994b; Tyler 2004). But this interest is based on the increasing applications of REEs as tracers of geochemical processes and most of the researches were carried out in natural waters, coal mines, and other mineral areas (Eskenazy 1999; Protano et al. 2002).
It is widely recognized that toxicity of heavy metals does not merely depend on the total concentrations but on their forms (Brummer 1986). Forms of REEs are directly related to their mobility, distribution, bioavailability, and ultimate fate in the environment (Ji et al. 2004). REEs are a group of elements which are relatively inert. As a result, REEs mainly exit in residue form in soils, sediments, and suspended particulates in water bodies in China with very low mobility.
Table 9 Concentrations of REEs in Dicraneopteris dicthotoma (micrograms per gram) Soil Mining areas
Non-mining areas
Root
Stem
Petiole
Lamina
Method
Reference
1
330.68
926.43
137.63
136.66
2,648.79
ICP-MS
Wang 2005
2
207.02
861.84
106.18
57.76
2.090.30
ICP-MS
Wang 2005
3
431.8
559.0
289.0
–
3,263.8
ICP-AES
Li et al. (1992)
4
806.4
955.9
287.0
–
2,825.8
ICP-AES
Li et al. (1992)
5
2,373
1,570
459
102
2,271
ICP-MS
Wei et al. (2001)
6
396
1,269
401
61
977
ICP-MS
Wei et al. (2001)
7
2,314
1,754
633
56
1,660
ICP-MS
Wei et al. (2001)
8
–
788.79
170.05
–
1,030.33
ICP-AES
Hong et al. (1999)
1
–
851.32
98.64
40.74
1,494.45
ICP-MS
Wang et al. (2010)
2
15.3
38.7
39.5
–
1,914.0
ICP-AES
Li et al. (1992)
3
111
134
107
51
1,121
ICP-MS
Wei et al. (2001)
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Environ Monit Assess (2014) 186:1499–1513 Atmosphere
b 3
Peroral intake a (average) 659.4µg/(d• pp)
3
Total Suspended Particulates 0.359µg/m (0.087µg/m ) 3 3 Inhalable Particles 0.613µg/m (0.328µg/m )
Atmospheric deposition (Dry and Wet)
Industrial Exhaust Emission
Inhalation Hair (workers)
a a
8.98µg/g (0.61µg/g)
Hair (farmers) 4.68µg/g (0.61µg/g)
Intake
Intake
Runoff Industrial drainage
b
Hyper-accumulate Plants d Dicranopteris linearis (leaves, dry mass) 2648.8µg/g 1121µg/g
Vegetables (dry mass) 6.37µg/g 0.878µg/g
Uptake
Soil b 863.73µg/g 240.47µg/g
Uptake
Well water b 9.18µg/L 0.38µg/L
c
Water body (DREE) 36.4µg/L (0.35µg/L) Irrigation
Leaching
Fig. 1 The average contents of REEs in Chinese REE mining area. a Lu et al. (1995), b Zhu et al. (2002), c Meng et al. (2007), d Wei et al. (2001)
However, in REE mining areas, the forms of REEs present significantly different features. Organically bound and exchangeable REEs are considered as the predominate forms in soils in REE mine areas of China (Gao et al. 2001; Yu et al. 2009; Zhang et al. 2012). Wei et al. (2001) determined content of soluble REEs in soils (extracted by 1.0 mol/L NaAc-HAc, pH=4.8) which could be used directly by plants in Jiangxi REE mining areas. They found that soluble REE concentrations accounted for 44–52 % of the total REE concentrations in the mining areas, whereas in non-mining area, the ratio was only 14 %. Although it has been found that residue form still accounted for the largest part of REEs in sediments and suspended particulates, noticeably increasing proportions of exchangeable REEs were observed (He et al. 2004a, b, c). All of those results indicated obviously higher mobility and bioavailability of REEs in Chinese REE mining areas. High mobility of REEs in the mining areas leads to high plant availability. Plants grown in REE mining areas, especially REE hyperaccumulators, contain a high level of REEs in their organs. D. dichotoma, one of REE hyperaccumulator and dominant species in
Chinese REE mining areas, was frequently used as ideal materials for studies on REE forms inside plants. But owing to the lack of sensitive instruments and practical analytical techniques which allow speciation analysis of REEs in biological materials, investigations on forms of REEs inside plants are still limited. Guo et al. (1996) separated two kinds of REE-bound proteins in leaves of D. dichotoma. Wang et al. (2001b) studied REE speciation in D. dichotoma (DD) by molecular activation analysis (MAA). The results showed that the REEs were firmly bound to the proteins, polysaccharides, and DNA in the leaves of DD. Zhang et al. (2000a) separated leaves of D. dichotoma collected from a rare earth ore area into ether fractions, water fractions, base fractions, and residues and determined the contents of REEs by instrumental neutron activation analysis (INAA). They found most of REs remained in the residues (pectin, hemi-cellulose, and cellulose). Distribution characteristics of REEs in Chinese mining areas The distribution of exogenous REEs in the environment of REE mining areas is an important aspect of
Environ Monit Assess (2014) 186:1499–1513
REE geochemical behavior. Under natural conditions, the translocation ability of REEs is relatively weak as compared with other metal ions. A vast majority of the exogenous REEs are fixed on solid surfaces and exist in inert forms, which are normally concentrated in the top layer of soils or deposited in sediments in water bodies (Jones 1997). However, areas around REE mining sites are usually overloaded with REEs. The forms of REEs varied with the change of environment. Studies revealed that REEs had high mobility in soils in Southern REE mining areas of China. The strong weathering and low pH value in soils are considered as the main reasons (Ma et al. 2004; Xu 2005). High proportions REEs have also been found in dissolved forms or absorbed by suspended particulates in the mining areas as compared to natural water bodies (He et al. 2004b; Meng et al. 2008), which also indicate a strong mobility of REEs in areas around mining sites. After uptaking by plants, REEs may enter different parts of plants. It has been observed that, under natural conditions, 80 % REEs are retained in roots and the concentrations of REEs in plant organs are always ranked in the order of root>stem>leaf>flower>fruit and seed (Ji et al. 2004; Yang 1996). Tagami and Uchida (2005) measured REE contents in plant organs cultured in REE-contained nutrient solution for 1 day by ICP-MS. They found that almost all of the absorbed REEs were distributed in fine roots and only small amounts of REEs were found in leaves. Feng et al. (2001) and Zhang et al. (2002) observed similar results in their researches. However, a totally opposite result has been found in hyperaccumulating plants grown in REE mining areas. REEs uptaken by REE hyperaccumulators in the mining areas are distributed mainly in their aboveground parts. Studies demonstrated the ∑REEs in D. dichotoma grown in the mining areas followed the order of leaf= root>stem>petiole (see Table 9). In addition, the age of leaves may also be a factor influencing REE distribution in plants. Xiao et al. (2003) determined REE contents in fern collected in South China by INAA, and the results showed that the older the leaves were, the higher REE concentrations they had. The distribution of REEs in plant organs at the cellular level is another hot topic. The limitation of analytical procedure has made it hard to locate REEs in plant cells. Some of the results are even contradictory, especially when they refer to whether lanthanides
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can pass through cell membranes. It is generally considered that REEs are fixed on the cell walls of plants mainly because of the strong absorbance by ionogens with negative charges on cell walls, such as carboxyl and hydroxyls (Han et al. 2005; Shan et al. 2003). In addition, the precipitation with phosphate in the cell gap could lead to a further difficulty for REE ions entering into cell membranes (Diatloff et al. 1993; Quiquampoix et al. 1990; Robert et al. 1996). However, with the development of the analytical methods in recent years, more and more results indicated that even though it is difficult, REEs ions may pass through cell membranes. Li et al. (1995) found that Ce could enter plant cells and especially be accumulated in cell nucleus determinated by EDX. Zhang et al. (2000b) determined REE contents in protoplasm of plants from REE mining areas by MAA and found REEs could enter cells. Gao et al. (2003) revealed that Eu and La existed in cytoplasm and were bond to chloroplast, mitochondria, and cell nucleus. Additionally, studies demonstrated that REEs could possibly be involved in plant growth activities. Hong et al. (1999) indicated that La could replace magnesium in the structure of chloroplast. Wei et al. (2000) demonstrated the same result in their study using EXAFS. Fractionation of REEs in Chinese mining areas It is known that all 15 REEs share similar properties, but they still have differences from each other. As a result, REE fractionation may take place during the transport processes. REE patterns in soils, suspended particulates, and sediments exhibit a light REE enrichment and negative Eu anomaly in REE mining areas of Northern China (He et al. 2004a; Li et al. 2008) and a heavy REE enrichment pattern with strong negative Eu anomaly in REE mining areas of Southern China (Gao et al. 2001; Yu et al. 2009). Results of these studies revealed that in soils and water bodies, the distribution patterns of REEs are all in line with their parent rocks. No obvious fractionation between REEs occurred during those transport processes. However, a significant fractionation phenomenon among individual members of REEs were observed during the transport processes from soil to plant, as well as the transportation inside plants (Miekeley et al. 1994; Shan et al. 2003). Previous studies had indicated several kinds of REE enrichment features in plant and
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the LREE-enrichment in the aboveground part was considered as the most common one in general. However, Ding et al. (2005b; 2007) observed heavy REE enrichment in leaves, light REE enrichment in roots, and middle REE enrichment in stems in their hydroponic experiment on Triticum aestivum L. and Glycine max L. Wang et al. (2005) found a light REE depletion in the aerial parts of Oryza sutiva. Furthermore, the tetrad effect caused by anomaly of individual REE elements was also demonstrated in several researches (Feng et al. 2002; Wyttenbach 1998). Several possible mechanisms of REE fractionation in plants have been discussed, including organic ligands bonding in rhizosphere, cell wall absorption, and P precipitation in cell gaps. Different mechanism may influence different stage of REE transportation from soil to various parts of plants. Liang et al. (2008) had a thorough discussion on the REE fractionation mechanisms in plants in their review and developed a conceptual model of fractionation of REEs in plants. Unlike characteristics of REE fractionation in plants, the light REE enrichment can be regarded as the representative feature of REEs in atmospheric particles whether in REE mining areas or not according to existing studies. La, Ce, Pr, Nd, Sm, Eu, and Y share almost 90 % of total atmospheric REE burden. The ratio of light REE to heavy REE and the ratio of light REE to total REE in atmospheric particles collected from different areas are presented in Table 10. The value of LREE/HREE in Chinese REE mining areas are 9.51 which is far above this ratio in local soils (2.26) (Gao et al. 2001). Influencing factors Same as under natural conditions, the behavior of REEs in REEs mining areas are influenced by many factors, such as their parent rocks, weathering processes, soil-
forming processes, climate, pH value, redox conditions, and organic matters (Cao et al. 2001; Heinz et al. 1993; Sholkovitz 1994b). Among these factors, pH value strikingly plays an important role in controlling translocation of exogenous REEs in the environment. Literature data suggested a strong negative significant correlation between pH value and REE mobility in the environment. The reason is that high pH value could significantly enhance the absorption of REEs, thus decreasing their mobility. Zhu et al. (1996) compared REE migration in soils under different pH conditions using numerical simulation and isotopic tracer. The results revealed that the migration distance turned out to be 1 cm for acid soils annually, 0.5 cm for slight acid soils annually, and no migration for the alkaline soils. Ran and Liu (1992) found increasing desorption of REEs from 29.5 to 90 % when pH decreased from 6.3 to 4.1. Zhu et al. (1993) also demonstrated that the adsorption of REEs was much greater in soils of higher pH due to the formation of complex ions (such as Ln(OH)2+, Ln(OH)2+, and Ln(OH)4− ) and their strong absorption at the sites covered with OH−. pH can influence REE behavior by controlling the presence and concentrations of metal ions in water body. The concentrations of REEs in water decrease significantly when the pH increases. The reason is the occurrence of the coagulation and precipitation of Fe–Al–Mn oxides/hydroxides which can effectively absorb and remove REEs from water. Oxidization and reduction conditions mainly control the behavior of Ce and Eu because of their changeable valence (Braun et al. 1990; Ma et al. 2004). Under normal conditions, all REEs exit as +3 ions except that Ce and Eu also occur as Ce4+ and Eu2+, respectively. Thus, Ce and Eu could be separated from the other REEs with the changes of redox conditions. Under oxidation conditions, Ce occur as Ce4+ which is relatively inert and could be stabilized by forming oxides or by adsorption onto solids, thus causing a significant positive Ce anomaly.
Table 10 The fractionation features of REEs in atmospheric particulate matters Non-mining areas
REE mining areas
Sampling location
Delft, Germany
Osaka, Japan
Beijing, China
Jiangxi, China
Jiangxi, China
LREE/HREE
11.06
19.83
11.47
14.11
9.51
LREE/∑REE
0.90
0.952
0.920
0.863
0.837
Reference
Wang et al. (2000a)
Sugimae (1980)
Wang et al. (2001a)
Zhu et al. 2003
Zhu et al. 2003
Environ Monit Assess (2014) 186:1499–1513
Conclusions After reviewing the collected data, we found that with the sharp increase of mining and producing activities, the environment of mining areas in China has dramatically changed in the past decades. Large amount of REEs released into the environment, scattered, and accumulated in soils, waters, atmosphere, and plants nearby through various transport processes. Results from various studies focusing on the contents of REEs in mining areas all showed a significantly high level of REE in every environmental component of mining areas, which was hundreds of times higher than non-mining areas. Furthermore, due to special mining environmental conditions, such as low pH, the behavior of REEs in mining areas presented differences from them in natural conditions. Comparing with natural conditions, REE existed in more active states in mining areas, such as exchangeable REEs in soils and dissolved REEs in water bodies. Those REEs were easier to move and transport among different environmental components, as well as to be intake by plants and enter into food chains which indicate possible harmful effect on human. Previous studies have already pointed out the accumulation of REEs inside local residents’ bodies which could be a threat to their health. However, studies on environmental behaviors of REEs in REE mining areas reviewed above were still limited. We identified some gaps in this research field which need more attention and further improvement. Firstly, owing to the more active states of REEs in mining areas and their harmful effects on human, a more systematic and in-depth research is needed on the translocation and circulation of REEs in various media in REEs mining areas. The identifications of each transport process of REEs and detailed analysis of their transfer characteristics, such as their forms and the amount of REEs in every single transport pathway, will provide us with a better understanding of the behaviors of REEs in mining areas and a strong base for further health risk assessment of REEs. In addition, with the increasing reports revealing the toxic effects of REEs, the harmful effects of REEs on human health is gradually recognized. People lived in mining areas could be exposed on REEs through different exposure pathways, such as inhalation from air, ingestion from food, drinking waters, and dermal absorption from soils. The high levels of REEs in soils, air, edible parts of plants, and well waters in Chinese REE mining areas are a great threat to human health. The significant
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inaccuracy of food REEs detection and the lack of certain standards make it difficult to evaluate health risks of REEs. Therefore, it is necessary to develop guideline values of REEs in foods, drinking waters, and all other environmental components. Secondly, as we know, high levels of toxic elements in the air have caused serious pollution and threatened human health. The concentrations of organic compounds and heavy metals, such as As, Cu, Pb, and Hg, in the air have received considerable attention in last decades. However, little work has been devoted to REEs in the air. By now, data about concentrations, forms, and distribution characteristics of atmospheric REEs in Chinese REE mining areas are inadequate. Considering the fact that inhalation is the main exposure pathway of REEs for people living around REEs mining areas, more studies focusing on the behavior of REEs in ambient air of REE mining areas in China should be carried out in the future. Furthermore, we suggest more attention to be paid on mining areas with windy conditions and smelting processes, both of which could contribute to further scattering of REES and lead to serious air pollution. At last, there are numerous methods for pretreatment and detection of REEs when measuring their contents and forms, such as spectrophotometric method, INAA, and NAA in the past decades, and more advanced methods like ICP-MS, ICP-AES, and EXAFS in recent years. However, the accuracy for different methods varies, and the detected number of REEs also varies. We suggest standardized methods on the analysis of REEs in the environment which will enhance the comparability for different researches. Acknowledgments This study was sponsored by the Key Project of the Knowledge Innovation Program of IGSNRR (2012ZD002), the Special Fund for the Public Interest Research of China MEP (no. 201209012), the National Key Project for Basic Research (no. 2012CB417004), and the National Natural Scientific Foundation of China (no. 41171390).
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