Potential Anthropogenic Pollution of High-technology ...

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Reviews 1

Potential Anthropogenic Pollution of High-technology 2

Metals with a Focus on Rare Earth Elements in 3

Environmental Water 4

5

Akihide ITOH,*, † Akane YAIDA,** and Yanbei ZHU*** 6

7

* Department of Environmental Science, School of Life and Environmental Science,8

Azabu University, 1-17-71 Fuchinobe, Chuo-ku, Sagamihara, Kanagawa 252-5201,9

Japan10

** Graduate School of Environmental Health Sciences, Azabu University, 1-17-71 11

Fuchinobe, Chuo-ku, Sagamihara, Kanagawa 252-5201, Japan 12

*** National Metrology Institute of Japan (NMIJ), National Institute of Advanced 13

Industrial Science and Technology (AIST), 1-1-1 Umezono, Tsukuba, Ibaraki 305-14

8563, Japan 15

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Contents 17

1. Introduction18

2. REEs in environmental water19

2.1 Analytical methods 20

2.2 REE concentration anomalies 21

2.3 Comparison of the degree of Gd anomalies 22

2.4 Chemical speciation of Gd compounds 23

2.5 Ecotoxicology and bioavailability of anthropogenic REEs 24

3. Multielement analysis of high technology metals as well as REEs25

4. Conclusion26

5. Acknowledgements27

6. References28

29

† To whom correspondence should be addressed. 30

E-mail: [email protected]

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Analytical SciencesAdvance Publication by J-STAGEReceived September 19, 2020; Accepted November 2, 2020; Published online on November 6, 2020DOI: 10.2116/analsci.20SAR16

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Abstract 33

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In recent years, utilization of high technology metals such as rare earth elements 35

(REEs), whose abundance are extremely low in the earth, has rapidly increased with 36

development of new types of industrial materials and pharmaceutical products. This 37

review overviews a new type of potential anthropogenic pollution of the high-technology 38

metals with a focus on REEs released in environmental water from a waste treatment 39

facility. In this paper, potential anthropogenic pollution was defined as pollution caused 40

by metals gradually enriched in the environment by human activity, although standard 41

and guideline concentrations of these elements are not regulated by environmental quality 42

standards for water pollution. We review the analytical methods of REEs and the potential 43

anthropogenic pollution of REEs with a focus on Gd in the viewpoints of comparison of 44

degree of Gd anomaly, chemical speciation, ecotoxicology, and bioaccessibility. 45

Moreover, we also highlight the comprehensive analysis based on multielement analysis 46

of high technology metals as well as REEs for the further screening of a potential 47

anthropogenic pollution. 48

49

50

Key words: ICP-MS; trace metal, REEs; Gadolinium, river water; coastal seawater; 51

ecotoxicology; bioaccessibility; speciation; concentration anomaly. 52

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1. Introduction 59

The presence of excessive heavy metals in the environment has negatively affected 60

the health of humans, animals, and plants. Hence, at present, in a general water quality 61

investigation of the environmental, the concentrations of many trace metals in 62

environmental water are controlled to monitor its quality as well as to protect human 63

health and/or aquatic life.1-3 For example, in Japan, Cd, Pb, Cr(VI), As, Hg (total and 64

alkyl-), Se, Ni, Mo, Sb, Mn, and U are regulated in the Environmental Quality Standards 65

for Water Pollution of Japan,1 while Al, Sb, As, Ba, Be, Cd, Cr (III and VI), Cu, Fe, Pb, 66

Mn, Hg (total and methyl-), Ni, Se, Ag, Tl, Sn (tributyltin) and Zinc are regulated in the 67

National Recommended Water Quality Criteria of the United States of America.2,3 Many 68

studies concerning the analysis of these elements have been reported.4-11 The 69

concentrations required for regulation are generally at mg L-1 to μg L-1 levels, and their 70

determination can be accomplished using inductively coupled plasma mass spectrometry 71

(ICP-MS), inductively coupled plasma optical emission spectrometry (ICP-OES), and 72

atomic absorption spectrometry (AAS). Various studies for heavy metals in 73

environmental waters such as rivers, lakes, and coastal seawater has been performed not 74

only for concentration,12-26 but also for speciation,27-29 environmental toxicology,30-34 and 75

assessment of aquatic ecosystem.35-37 76

In recent years, utilization of high technology metals such as rare earth elements 77

(REEs), whose abundance are extremely low in the earth, has rapidly increased with 78

development of new types of industrial materials and pharmaceutical products.38,39 In 79

Japan, the Ministry of Economy, Trade and Industry, defined these elements as “rare 80

metal”. 40 These include 31 mineral species and 47 elements. If materials including these 81

elements were disposed without appropriate control, a new type of environmental 82


4

pollution may occur. Thus, we became interested in the research for a new type of 83

potential anthropogenic pollution of high-technology metals with a focus on REEs in 84

environmental water. In this paper, potential anthropogenic pollution was defined as 85

pollution caused by metals gradually enriched in the environment by human activity, 86

although standard and guideline concentrations of these elements are not regulated by 87

environmental quality standards for water pollution. In particular, many researches for 88

anthropogenic REEs in environmental water have been reported worldwide. The 89

significant increase in the use of REEs in high technology products and processes has led 90

to an increased release of these metals into the environment. Hence, the determination of 91

REEs in the environment have been performed not only from the viewpoint of 92

geochemistry and geoscience, but also from the environmental concern. The first report 93

of the anthropogenic REEs in the environmental water was published in 1996,41 when the 94

concentration anomaly of Gd was identified in the Havel River in Germany. Nowadays, 95

it is widely known that the relative abundance of Gd in the river water and coastal 96

seawater around large metropolitan areas is typically higher than those of its neighboring 97

REEs, i.e. Eu and Tb.15,39,41-64 This was artificially caused by anthropogenic sources 98

mainly due to use of Gd compounds used as a contrast reagent for magnetic resonance 99

imaging (MRI) in medical diagnosis. 100

In this paper, we review the analytical methods of REEs and the potential 101

anthropogenic pollution of REEs with a focus on Gd in the viewpoints of comparison of 102

degree of Gd anomaly, chemical speciation, ecotoxicology, and bioaccessibility. 103

Moreover, we also highlight the comprehensive analysis based on multielement analysis 104

of high technology metals as well as REEs for the further screening of a potential 105

anthropogenic pollution. 106


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107

2. REEs in environmental water 108

2.1 Analytical methods 109

Nowadays, ICP-MS is the dominating approach for the determination of REEs in 110

environmental water.43, 49, 50, 65-96 Other works reported the determination of REEs using 111

neutron activation analysis (NAA)97,98 and thermal ionization mass spectrometry 112

(TIMS).51,99,100 The increasing application of ICP-MS in the determination of REEs can 113

be attributed to increasing availability of ICP-MS in the laboratories for analysis, as well 114

as to the capability of ICP-MS for simultaneous determination of REEs at pg mL-1 level. 115

A comparison of analytical performance of ICP-MS-based techniques for 116

determination of REEs in environmental water is summarized in Table 1. As can be seen 117

from Table 1, the lowest value of detection limit obtained by direct measurement with 118

ICP-MS varied from 0.003 pg mL-1 for Tm to 1.9 pg mL-1 for Dy.49,75,76, 78, 85 The detection 119

limits would depend both on the model of ICP-MS instrument and on the operating 120

conditions. A sample volume not more than 10 mL can be processed for the determination 121

of REEs by ICP-MS when a preconcentration factor under 10-fold is enough.49,77 A 122

preconcentration factor more than 30-fold are usually obtained with a sample volume 123

from 40 to 5000 mL.66,-69, 80, 81, 90, 94-96 The lowest value of detection limit reported for 124

REEs determined by ICP-MS varied from 0.001 pg mL-1 to approximately 2 pg mL-1. The 125

reported analytical performances in Table 1 might indicate that the lowest value of 126

detection limit for measurement of REEs in environmental water by ICP-MS with 127

preconcentration depends on both the operating conditions and the pretreatment 128

techniques.49, 65- 71, 73, 75-78, 80, 81, 85 90, 94-96 129

In order to evaluate the anomaly of some REEs, scientists often attempt to determine 130

the whole group of REEs instead of selected ones that can be determined directly with 131


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ICP-MS. However, in this case, some REEs are often near or below the detection limit 132

of an ICP-MS instrument. Consequently, preconcentration of REEs must be carried out 133

prior to the ICP-MS measurement. In addition, removal of high contents of salt (ca. 3.5 %) 134

is often required for measuring REEs in seawater samples so as to reduce the burden to 135

the introduction system (e.g., torch, cones, and ion-lens, etc.) of an ICP-MS instrument. 136

There are multiple approaches reported for the preconcentration of REEs prior to the 137

measurement, including coprecipitation with Fe(OH)3,51, 65, 95, 100 coprecipitation with 138

Mg(OH)2,81 solid phase extraction (SPE) with chelating resin, 67, 69, 70, 72, 80, 87, 89, 93, 96-97 139

cation exchange,71 chromatography,73,74 solvent extraction,78 magnetic SPE with 140

nanoparticles,90 and extraction with calcium alginate microparticles.94 141

The majority of studies employed the SPE with chelating resins for 142

preconcentration of REEs prior to the measurement by ICP-MS. It can be attributed to 143

the merits of the good selection for REEs without introduction of a new metal-matrix as 144

well as the relatively simpler operations than cation exchange. Laboratory-modified C18 145

cartridge, on which the complexing agent composed of a mixture of bis(2-ethylhexyl) 146

dihydrogen phosphate (HDEHP) and 2-ethylhexyl dihydrogen phosphate (H2MEHP) was 147

loaded, was also reported as an effective SPE for preconcentration of REEs in seawater 148

samples.66 This preconcentration method was either performed by utilizing the complex 149

formation between REEs and chelating resin or reagent. They have often been applied to 150

the determination of all REEs in environmental water including seawater by many 151

research groups to investigate the concentration anomalies.41, 42, 44-46, 48 However, due to 152

increasing availability of commercial chelating resins, in recent years, they become a 153

popular choice for SPE processes and concentration of REEs. 154

It is remarkable that the presence of organic compounds (e.g., 155


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ethylenediaminetriacetic acid at nmol L-1 level) in environmental water samples affects 156

the recovery of REEs during the preconcentration process involving the complex 157

formation of REEs,47, 96 despite acidification of filtered water samples to pH 1 by the 158

addition of HNO3 or HCl. The reason for this phenomenon can be attributed to the 159

competition between refractory organic anions and functional groups of chelating resin 160

(e.g., carboxyl group) and reagents in adsorbing of REEs in environmental water samples. 161

Hence, decomposition of organic compounds by acid treatment with HNO3 and H2O2 at 162

170 ℃ for 4 h was proposed prior to the SPE operation with chelating resin to improve 163

the recovery of REEs.96 164

Polyatomic spectral interferences must also be considered for accurate determination 165

of REEs by ICP-MS. Typical polyatomic spectral interferences are summarized in Table 166

S1. Typical polyatomic spectral interferences and the resolution required for separation 167

in Table S1 are estimated from the isotopic mass of each element.102 In most of the 168

reported works, corrections of such polyatomic spectral interferences were carried out 169

based on mathematic calculations of the production rates of these interfering species. It 170

is notable that the removal of Ba from the sample during the preconcentration process for 171

REEs helps to reduce spectral polyatomic interferences with the measurement of 139La 172

(138Ba1H) and 151Eu(135Ba16O), and 153Eu (137Ba16O). This is especially important because 173

the concentration of Ba in environmental water samples is usually much higher than those 174

of La and Eu by three-to-four orders of magnitudes. Thus, spectral interferences can be 175

problematic, even if the ICP-MS measurement is performed under beneficial conditions 176

of very low ratio of BaH+/Ba+ and BaO+/Ba+. 177

Fortunately, the availability of tandem quadrupole ICP-MS with a reaction cell (also 178

known as ICP-QQQ) permitted the complete separation of polyatomic spectral 179


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interferences from isotopes of REEs by using oxygen as the reaction gas and measuring 180

the REEs as their mono-oxide ions.103 In the near future, there may be more applications 181

by this type of ICP-MS instrument to the determination of REEs in environmental water. 182

The precision for the results of REEs might depend on the instrument for 183

determination, preconcentration technique, and preconcentration factor. When the 184

concentrations of REEs in the measured solutions are sufficient for determination, a 185

typical relative standard deviation around 1% to 5% can be achieved.66, 76, 96 Such 186

precision is no problem for the evaluation of anomalies of REEs in natural water samples, 187

because the concentration anomalies in the REE patterns are often illustrated as logarithm 188

scale. Isotope dilution analysis had also been reported for the determination of REEs in 189

natural water samples and contributed to improving the accuracy of the results with 190

precisions comparable to those obtained by calibration curve methods.104, 105 The 191

application of multi-collector (MC-) ICP-MS to isotope dilution analysis of REEs might 192

help to improve the precision of measurement. However, there is few reports about 193

concentrations of REEs in natural water obtained by isotope dilution MC-ICP-MS, 194

Moreover, it is difficult to investigate concentration anomalies of REEs using the isotope 195

dilution-ICP-MS, because mononuclidic REEs such as Pr, Tb, Ho, and Tm can not be 196

determined. Procedure blank values for the determination of REEs in natural water are 197

usually controlled at a sufficiently lower level in comparison to those in the samples. 198

199

2.2 Concentration anomaly of REEs 200

It is widely known that the concentration distributions of REEs in environmental 201

water typically show zig-zag patterns, according to the Oddo-Harkins rule.106 This 202

indicates that the concentration of REEs with even atomic numbers are higher than those 203

of the neighboring REEs with odd atomic numbers. Thus, REE patterns, normalized by 204


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the concentrations of REEs in Post Archean Average Australian Shale (PAAS),107 or North 205

American Shale Composite (NASC)108 are often used to identify the occurrence of a 206

concentration anomaly of REEs by comparing the relative abundance of REEs in 207

environmental water. Generally, it is considered that PAAS is the most widely used 208

composite, as it represents the most appropriate average earth crust. The REE patterns for 209

some environmental waters, which were normalized using PAAS, are shown in Fig. 1.48, 210

109 The REE patterns with the positive anomalies of anthropogenic Gd and with those of 211

La and Sm as well as Gd are shown in Figs. 1B48 and 1C,109 respectively, while the REE 212

patterns for pristine river waters in clean area without anthropogenic contamination are 213

shown in Fig. 1 A. As can be seen in Fig.1 A,48 the REE patterns for the pristine river 214

water showed relatively smooth curve after normalization. In these patterns, heavy REEs 215

provide relatively higher normalized values than light ones and no anomaly except for Ce. 216

The different extents of relatively enrichment of heavy REEs might be attributed to some 217

possible factors that stronger complexing ability of heavy REEs with dissolved organic 218

matters than those of light ones14, 110 and removal of colloidal light ones during filtration 219

process.111 It should be noticed here that the negative anomalies of Ce were observed. 220

This is a typical REE pattern for natural water without anthropogenic contamination. 221

REEs usually exist in the oxidation state of +3 with quite similar ionic radii, so that they 222

generally provide the similar chemical properties and behavior. However, Ce (III) is easily 223

oxidized to Ce (IV) in natural water under aerobic conditions, and precipitated as 224

Ce(OH)4 and removed from water.43,111 Thus, the negative anomaly of Ce is often 225

observed for natural water. 226

On the other hand, in Fig. 1B, the positive anomaly of Gd was clearly observed in 227

the REE patterns for environmental water after mixing with discharge water from waste 228


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treatment facility (WWTF). It is well known that this was caused by environmental 229

outflow of Gd compounds used as contrast reagent for MRI in medical diagnosis, 230

although Gd is also used as the manufacture and polishing of glass products, the 231

phosphors in computer monitor, and color television tubes and the nuclear reactor control 232

rod etc. The Gd compounds are administrated to patients in a hospital and are then 233

excreted as waste and transported into environmental water via WWTP. They are directly 234

released into river water from the discharging water without eliminating during the waste 235

treatment process in WWTF. Consequently, Gd anomalies were observed in 236

environmental water around large metropolitan area. The concentration anomalies due to 237

the anthropogenic Gd have been reported for all kind of environmental waters including 238

river waters, lake waters, groundwater, and coastal seawater all over the world since the 239

mid-1990s.41-64 In addition, the Gd anomalies in the REE patterns were observed in tap 240

water sampled in Berlin along the Havel River and other German cities along the Rhine 241

River as well as in the environmental water.107 The anthropogenic Gd was also found in 242

tap water and tap water-based beverages from fast-food franchises in six major Germany 243

cities.108 244

In recent years, the positive concentration anomalies of La, Sm, Ce, Eu as well as Gd 245

have been reported.109, 112 The occurrence of a positive Ce anomaly is influenced by its 246

redox state in the environmental water; therefore, it is not necessarily caused by 247

anthropogenic micro-contamination. However, other anomalies for REEs would be 248

caused by anthropogenic micro-contaminant. As shown in Fig. 1C, Kulaksiz and Bau109 249

reported that anthropogenic Sm and La as well as anthropogenic Gd was detected in the 250

Rhine River near the border between Germany and the Netherlands. Based on analysis 251

using ultrafiltration, it is suggested that while the anthropogenic Gd is not particle-252


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reactive and thus exclusively present in the truly dissolved REE fraction (<10 kDa), the 253

anthropogenic La and Sm are present in the colloidal/nanoparticulate REE fraction (10 254

kDa-0.2 m). Although the origin of Sm was unclear, it is considered that La was derived 255

from a point source where industrial cracking catalysts for petroleum refining were 256

produced. However, there is no obvious reason why chemical speciation of Sm should 257

differ from that of anthropogenic La. Their analytical results indicate that anthropogenic 258

Sm might originate from the same industrial cracking catalyst production effluent that 259

causes the La positive anomaly. Sm compounds are utilized in a high-strength permanent 260

magnet to control rods in nuclear reactors. They are also employed as catalysts in the 261

decomposition of plastics, dichlorination of polychlorinated biphenyls (PCBs), and 262

dehydration and dehydrogenation of ethanol. Furthermore, SmI2 is commonly used as 263

reducing and coupling reagents in organic synthesis. On the other hand, as it can be seen 264

from Fig. 1 C, positive anomaly of Pr as well as Gd, La, and Sm was observed in the 265

samples form the industrial cracking catalyst production effluent pipe at Rhine River. 266

However, this was not mentioned in the report by Kulaksiz and Bau.109 267

The Eu anomaly can be caused by the presence of Eu-enriched minerals in the 268

vicinity of environmental water and/or the result of the unique redox behavior. 269

Specifically, Eu3+ can be reduced to Eu2+, which exhibits increased solubility. The 270

reduction of this element occurs more readily than it does in the case of other REEs. Thus, 271

the anomaly of Eu due to naturally-occurring Eu were often observed in the samples from 272

the upstream river water to estuarine water.113 On the other hand, when the river water 273

samples are measured in the quadrupole ICP-MS, 151Eu and 153Eu is often interfered with 274

135Ba16O and 137Ba16O formed at the relatively high concentration of Ba in the samples. 275

Hence, in the case of direct ICP-MS measurement, the Eu anomaly has sometimes been 276


12

observed in the REE patterns as a result of spectral interference. In contrast, Itoh et al. 277

reported that the concentration anomaly of anthropogenic Eu in the Sakai-River was 278

caused by the effluent of a specific WWTF, even after the removal of a large proportion 279

of Ba by SPE with chelating resin and correction by spectral interference coefficient 280

method.112 Nowadays, the main origin of Eu is unknown, although Eu is used in liquid 281

crystal displays, fluorescent lighting, glass additives etc. 282

283

2.3 Comparison of the degree of Gd anomalies 284

To compare the Gd concentrations in samples from various environmental waters 285

around the world, it is important that the degree of the positive anthropogenic Gd 286

anomaly is objectively evaluated. However, when such comparisons are made, it should 287

be considered that the Gd concentrations are influenced by rainfall or naturally-288

occurring REEs in rocks and soils in the vicinity of the river water. Thus, a conversion 289

of the positive Gd anomaly into numerals is performed for each REE pattern using two 290

indexes. One index is the Panomaly (%), which indicates the percentages of the 291

anthropogenic Gd to the naturally-occurring one. It was defined by Zhu et al.43 and Yaida 292

et al.47 using the following equation (1): 293

𝑃anomaly = GdSN−GdSN∗

GdSN∗× 100(%) (1) 294

where GdSN is the shale (PAAS)-normalized concentration of Gd observed in 295

environmental water, and GdSN* is the theoretical shale-normalized concentration of Gd 296

when the positive anomaly is not observed in the REE patterns. GdSN* is approximated 297

from the straight line obtained from PAAS-normalized concentrations of Sm, Tb, and Dy 298

using the least square method. When Panomaly shows higher values (i.e., > 100%), it 299

suggested that the contribution rate of anthropogenic Gd to naturally occurring Gd is 300


13

larger. 301

Another index is GdSN/GdSN*, which was defined by the first reports for the positive 302

Gd anomalies in river waters by Bau and Dulski using the following equation (2) 303

GdSN/Gd*SN = GdSN/(0.33Sm SN + 0.67TbSN) (2) 304

However, Gd*SN was often estimated by employing other extrapolation methods. For 305

example, Kulaksiz and Bau proposed that Gd*SN was extrapolated using the shale-306

normalized values of Pr, Nd, and Sm in the REE pattern. In any case, a value of 307

GdSN/Gd*SN larger than 1.0 indicates the presence of an anthropogenic Gd in the 308

environmental water. 309

The Gd concentration and GdSN/Gd*SN index for natural land water, such as river 310

water, and for seawater samples obtained in various parts of the world are summarized in 311

Tables 2 and 3, respectively. The values of GdSN/Gd*SN were estimated from the 312

previously reported analytical data. Gd*SN was extrapolated using the shale-normalized 313

values of Sm, Tb, and Dy. It is noteworthy that the Panomaly and GdSN/Gd*SN index are 314

similar. The former is calculated by subtracting 1 from the value of GdSN/Gd*SN and 315

multiplying by 100. GdSN/Gd*SN was employed as the index in Tables 2 and 3, in which 316

the highest Gd concentrations were shown when several Gd concentrations were reported 317

for the same environmental waters. The GdSN/Gd*SN index values is considered to show 318

objectively the degree of Gd anomalies to some extent, while they are governed by the 319

mixing ratio of the river water and the discharge water from WWTF. In Tables 2 and 3, 320

the data for each natural land water and seawater are arranged in descending order of the 321

GdSN/Gd*SN index values. 322

As can be seen in Table 2, many studies on the Gd anomaly in natural land waters 323

such as river waters, have been reported so far. The reported concentrations of Gd ranged 324


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from 1 ng L-1 to several hundred ng L-1. In addition, the GdSN/Gd*SN index ranged from 325

1.1 to more than 2.0 in large metropolitan areas. This indicates that some parts of Gd 326

dissolved in natural land water have the anthropogenic origin, even in river water in the 327

suburb area. The Gd concentration and the GdSN/Gd*SN index in Havel River (Germany) 328

in 2009 were 492 ng L-1 and 653, respectively. Notably, these were the highest values 329

among all reported for natural land waters. This data indicates that the concentration of 330

anthropogenic Gd were more than 650-fold higher than the concentration of naturally-331

occurring Gd in the Haver River in 2009. Especially, larger values of GdSN/Gd*SN over 332

10 were seen in the river water samples in Germany, Japan, Brazil, and Poland. Most of 333

these reports was published after 2010. Incidentally, the owned number of MRI was in 334

order of the United States, Japan, and Germany. 335

On the other hand, the concentrations and the index values for the seawater were shown 336

in Table 3. As can be seen in Table 3, the Gd concentrations for seawater are generally 337

lower than those for natural land water, even in coastal seawater. This is because their Gd 338

concentrations were diluted by large amounts of seawater. The Gd concentration and 339

GdSN/Gd*SN index value were the highest for the coastal seawater in Tokyo Bay in 2008 340

among all reports. The Gd concentration and GdSN/Gd*SN index obtained from the same 341

sampling sites reported more than twice in different years are summarized in Table 4. As 342

can be seen in Table 4, for the Havel, Rhine, Elbe, and Tama Rivers as well as for the San 343

Francisco and Tokyo Bays, the values of the Gd concentration and GdSN/Gd*SN obtained 344

in recent years are larger than those determined in the past. Hence, it is suggested that 345

degree of Gd anomaly has been increasing annually in most of environmental water. 346

It is considered that the ecotoxicological effects would be not so high because the Gd 347

concentration level is very low under 1 g L-1, even in the Havel River showing the 348


15

highest value. Nevertheless, considering that the concentrations of anthropogenic Gd 349

have been gradually increasing in many river waters and coastal seawaters, environmental 350

monitoring must be performed to assess the risk to the aquatic ecosystem and human 351

health. 352

353

2.4 Chemical speciation of Gd compounds 354

It is important to elucidate the chemical species of Gd as well as the total Gd 355

concentration to investigate the ecotoxicological influences of the anthropogenic Gd on 356

the aquatic animals and plants and to assess its risk to human health. To evaluate the 357

chemical speciation of Gd in environmental water, the separation of Gd compounds used 358

as the contrast agents were attempted employing various analytical methods. However, 359

this process is extremely challenging due to the low concentrations of such species. The 360

main Gd compounds used as the contrast agent of MRI, which are all Gd complexes, are 361

summarized in Fig. 2.114 362

The hyphenation of separation techniques such as HPLC with ICP-MS have been 363

employed for speciation analysis of Gd in the environmental water.114-121 The ion 364

chromatography (IC)115 and the size exclusion chromatography (SEC)116, 117 were used as 365

separation method to separate Gd compounds used as the contrast agents and the Gd3+, 366

which is the most toxic species. In the IC/ICP-MS, Gd-DTPA (Magnevist) and Gd3+ could 367

be separated using the cation exchange separation column. The SEC/ICP-MS enabled the 368

separation of several Gd complexes and Gd3+ to some extent. Nevertheless, the 369

differences in the sizes of Gd complexes are too small for efficient SEC separation. On 370

the other hand, a hyphenation of hydrophilic interaction liquid chromatography (HILIC) 371

and ICP-MS has been employed as highly sensitive method for speciation analysis of 372

several Gd-based contrast agents in hospital effluents, river waters, WWTP effluents and 373


16

tap water. Birka et al. reported the sensitive quantification of the Gd-based contrast 374

reagents in river water samples using HILIC/ICP-MS.118-120 In their studies, this river 375

water sample was collected from a nature reserve in the city of Münster in Germany, 376

where the effluent from the city’s main WWTP enters the environment. Three kinds of 377

contrast agents, namely Gd-DTPA, Gd-DOTA (Dotarem) and Gd-BT-DO3A (Gadovist), 378

were identified and quantified in the obtained samples. The concentrations ranged from 379

0.59 nmol L-1 for Gd-DOTA up to 3.55 nmol L-1 for Gd-BT-DO3A. As a result of mass 380

balancing, the concentration of the Gd-based contrast agents was found to account for 74-381

89% of the total Gd concentration, thus indicating the presence of other Gd species. 382

Lindner et al. used a new zwitterionic HILIC column (ZIC-cHILIC) for speciation of Gd-383

containing contrast agents in tap water.120 ZIC-cHILIC contained a phosphorylcholine 384

moiety instead of a sulfobetaine group on the ZIC-HILIC. The positively charged head 385

group of the phosphorylcholine functionality in ZIC-cHILIC produced more surface 386

charges, which could interact with the negatively charged contrast agents more efficiently. 387

Overall, the more effective separation of negatively charged Gd-based contrast agents 388

from the new stationary phase was performed using ZIC-cHILIC. ZIC-cHILIC-ICP-MS 389

makes it possible to separate and determine the five contrast agents, i.e., Gd-BOPTA 390

(Multhiance), GD-DTPA, Gd-DOTA, Gd-DTPA-BMA (Omniscan), and Gd-BT-DO3A. 391

The chromatogram for Gd compounds in tap water samples obtained in Berlin (Germany) 392

measured by ZIC-cHILIC-ICP-MS are illustrated in Fig. 3. As shown in Fig. 3, three Gd 393

species, specifically Gd-BT-DO3A, Gd-DOTA and Gd-BOPTA, were found in tap water 394

samples at concentrations of approximately 10-20 ng L-1. It is noteworthy that these were 395

the same Gd species that had been previously detected in the surface waters, such as rivers 396

and lakes, in the western part of Berlin.45 This implies that the contrast agents were 397


17

transported from the surface waters through the bank filtration water into the tap; 398

therefore, it is possible that they entered the food chain. 399

400

2.5 Ecotoxicology and bioaccessibility of anthropogenic REEs 401

As described in Section 2.3, recent studies have showed that the concentrations of 402

Gd and some other REEs levels continue to increase in the environmental water and/or in 403

tap water. Nonetheless, their ecotoxicological effects as well as their influence on the 404

aquatic life and human health remain poorly understood. REEs are considered of minor 405

environmental concern due to their low toxicity to mammals.34 Although REEs have 406

received attention as environmental micro-contaminant in recent years, little is known 407

about their influences on aquatic organisms. Despite being considered as safe for many 408

years, in 2006, it was found that administration of Gd-based contrast agents led to serious 409

side effects, nephrogenic systemic fibrosis (NSF).122 It is speculated that the toxicity of 410

Gd-based contrast agents was predominantly caused by Gd3+ dissociated from them. The 411

ionic radius of Gd3+ is roughly equal to Ca2+, resulting in the blockage of the calcium ion 412

channels in the cells.64 Moreover, Gd3+ regularly binds with a higher affinity than Ca2+ 413

and completes with it during physiological processes.39 Hence, it is essential to further 414

investigate the bioaccessibility and ecotoxicological impact of anthropogenic REEs. 415

Studies on bioaccumulation and bioaccessibility of REEs have been performed mainly 416

using fishes, bivalves and algae.34 In this review, we introduce the studies using bivalves 417

for assessment of the influences of REEs on aquatic organism. 418

Merschel and Bau examined the bioaccessibility of anthropogenic La, Sm, and Gd in 419

river water by analyzing the aragonitic shell of freshwater mussel Corbicula fluminea.123 420

The shale (PAAS)- normalized REE patterns for Corbicula fluminea is illustrated in Fig. 421


18

4. The results in Fig. 4B indicate that all shells sampled at sites downstream of the 422

industrial point source of anthropogenic La and Sm in the Rhine River (at Worms) show 423

the positive La and Sm anomalies (the industrial point source of anthropogenic Sm was 424

situated only at Leverkusen). On the other hand, anomalous enrichment of Gd was not 425

observed in any of the analyzed shells. Hence, it was suggested that anthropogenic La 426

and Sm, which were detected in tap water as well as in river water, are bioaccessible. In 427

contrast, the anthropogenic Gd was not incorporated into the shells. The outcomes 428

regarding the incorporation of anthropogenic Gd in the Corbicula fluminea shells were 429

consistent with previous evidence concerning the long environmental half-life of Gd-430

based compounds used as contrast agents. Additionally, the results of this study further 431

implied that the conservative behavior of anthropogenic Gd makes it a useful tracer of 432

WWTP effluent mixed in river, lake, ground and tap waters. It is noticeable that the PAAS 433

normalized REE patterns of shells showed apparently relative depletion of heavy REEs 434

in comparison to those of water samples plotted in Fig. 1. This characteristic of PAAS 435

normalized REE patterns of shells might reflect the initial components of REEs in 436

particulates taken by the Corbicula fluminea or more favorable intake of heavy REEs 437

than light ones during the shell formation process. 438

On the other hand, Perrat et al. investigated the bioaccumulation of Gd in the tissues 439

(the digestive gland and the gill) in two kinds of freshwater bivalves (Dressena 440

rostriformis bugenisis and Corbicula fluminea) by in situ and in laboratory experiment.124 441

The data obtained in the laboratory experiment suggest that the bioaccumulation of Gd in 442

the digestive grand of the bivalves was observed when they were exposed to 1 and 10 443

g/L of Gd-DOTA during 7 and 21 days. These results demonstrated that Gd could 444

bioaccumulate in bivalve tissues even in the form of Gd-based contrast agents. In the 445


19

same study, it is confirmed that biomarkers such as glutathione-S-transferase, lipid 446

peroxidation, and electron transport system in C. fluminea were disturbed by the exposure 447

of Gd-DOTA. However, a return to the basic activities was observed within 21 days of 448

exposure. This indicates that C. fluminea was acclimated to the presence of Gd-DOTA 449

within 21 days. Moreover, Parant et al. studied the impact of Gd-based contrast agents 450

(Gd-DOTA and Gd-DTA-BMA) on the growth of the Zebra fish cell line (ZF4; ATCC 451

GRL-2050) under environmental concentrations.125 The toxicity of Gd-DOTA was not 452

observed; however, it was measured that the cell growth was decelerated. The same effect 453

was noted for a different fish cell line and another contrast agent (GD-DTPA-BMA). 454

Nevertheless, it is not clear whether the slower cell growth was caused by the Gd3+ or the 455

chelating structure of the contrast agent. 456

The above-mentioned studies about bivalves indicate that the Gd-based complexes 457

used for MRI are bioaccessibility and can be observed in the tissues of the organism. 458

However, such intake of Gd-based complexes might not result in the generation of Gd3+ 459

in the tissue of bivalves and did not lead to accumulation of Gd in the bivalve shells. The 460

intake of Gd-based complexes in the tissues of bivalves suggests a relatively shorter 461

biological half-life of the compounds in comparison to the natural-occurring Gd3+, which 462

compete with Ca2+ and can be accumulated in the bivalve shells in spite of extremely low 463

concentration. 464

465

3. Multielement analysis of high technology metals as well as REEs 466

In recent years, in addition to REEs, platinum group elements (PGEs), mainly Pt, Pd, 467

and Rh) has been increasingly reported in water environment.126-134 The increase of PGEs 468

in the environment can be attributed to the increasing use of these elements in the modern 469


20

industries related to autocatalytic converters, medical devices, and solar energy.135-140 In 470

addition, metal product processing of Ni and Cr smelters as well as medical applications 471

of Pt-containing drugs (e.g., cisplatin and carboplatin) have been considered as sources 472

of PGEs.140 There is currently no evidence regarding the discharge of PGEs by solar 473

energy products. However, the consumption of PGEs in solar energy industry implies that 474

it can be another source for PGEs discharge to the environment.140 There is also indication 475

of toxicity of PGEs to animals depending on the chemical species and the concentration 476

of the elements.129 Hence, further studies on the determination of PGEs in the water 477

environment are required. 478

The concentrations of PGEs in the industrial discharge water could be tens of ng L-479

1.126 The specific levels of the elements can be determined directly with analytical 480

instruments such as ICP-MS. However, spectral interference is one of the important issues 481

in the measurement of PGEs by ICP-MS. SPE with an alumina column was reported as 482

an effective approach for separation from spectral interferences as well as for the 483

preconcentration of PGEs prior to analysis by ICP-MS. A nominal enrichment factor of 484

PGEs up to 100-fold was achieved using only 30 mL of water sample and provided a 485

detection limit of 1 ng L-1.129 Furthermore, dispersive liquid-liquid microextraction 486

providing a nominal enrichment factor of 27- to 75-fold with 35 mL of water sample was 487

also reported for the measurement of PGEs by ICP-MS.131 It is notable that typical 488

concentrations of PGEs in unpolluted environment waters are less than 1 ng L-1.126 489

Quantitative analysis of such small amounts is challenging, even with an ICP-MS 490

instrument after preconcentration with a factor of 100-fold. This fact may be the major 491

reason for that PGEs in unpolluted environmental water have been rarely reported so far. 492

Several studies on the comprehensive determination of trace metals including high-493


21

technology metals as well as REEs were also performed for the screening of a potential 494

anthropogenic pollution in addition to the geochemical concern.15, 141-144 It is remarkable 495

that Vriens et. al reported a national survey of 69 element fluxes in wastewaters from 64 496

municipal WWTP and river waters in Switzerland.141 They showed average per capita 497

fluxes of each elements discharged from WWTP on addition to the spatial distribution of 498

many individual elements. The average per capita (population-weighted) fluxes of 62 499

elements discharged by wastewater from WWTF are shown in Fig. 5. Per-capita element 500

fluxes ranged from <10 g day-1 (e.g., for Au, In, and Lu) to1mg day-1 (e.g. for Zn, Sc, Y, 501

Nb, and Gd) and > 1 g day-1 (e.g. for P, Fe, and S). Effluent loads of some elements 502

contributed significantly to riverine budgets (e.g. 24% for Zn, 50% for P, and 83% for 503

Gd), indicating large anthropogenic inputs via the wastewater stream. 504

In recent decades, the development and availability of high-performance multi-505

elemental measurement instruments, particularly ICP-MS, has enabled the analysis of all 506

natural elements in the periodic table. Meanwhile, the industrial and medical application 507

of increasing amounts of elements may lead to anthropogenic pollution of the natural 508

water environment. Thus, comprehensive monitoring of almost all natural elements in the 509

periodic table, including high-technology metals, is required to trace and evaluate the 510

potential input of anthropogenic pollution from WWTF. 511

512

4. Conclusion513

In this review, the anthropogenic Gd outflowed into the environmental water was514

overviewed from various points of view. Studies on these potential anthropogenic 515

pollutions of Gd originated from MRI contract agents have been increasing annually 516

worldwide. They are also important as one of the PPCPs (Pharmaceutical and Personal 517


22

Care Products) pollutions for inorganic medical agents. In Germany, it is confirmed that 518

these Gd-based contrast agents were mixed in the tap waters of main cities as well as in 519

river waters. On the other hand, considering that the concentrations of anthropogenic Gd 520

have been gradually increasing in some river water and costal seawater near large 521

metropolitan area in Germany, Japan, and the United State, environmental monitoring 522

must be performed to assess the risk to the aquatic ecosystem. As the human health risk 523

of anthropogenic Gd as well as the ecotoxicology and the bioavailability were still unclear, 524

they have to be elucidated with chemical speciation and environmental dynamics of 525

anthropogenic Gd in near future. Moreover, the potential anthropogenic pollution of other 526

high technology metals in addition to Gd and other REEs in environmental water should 527

be surveyed to prevent the expansion of environmental pollution. 528

529

5. Acknowledgements530

The present work was supported by JSPS KAKENHI Grant Numbers JP19K12300.531

532

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755


Table 1 A comparison of analytical performance of ICP-MS based techniques for determination of REEs in natural water samples

Samples Pretreatment ICP-MS Va)/ mL PFb) Lowest value of detection

limit / pg mL-1 Year Reference

Groundwater Fe(OH)3 coprecipitation VG 12-12S NAc) NAc) 1 for Pr, Tb, Ho, Tm, Lu 1992 65

Seawater SPE with modified C18 resin VG Plasma Quad 1000, 5000 200-1000 NAc) 1992 66

Seawater SPE with chelating resin SPQ-8000 1000 100 0.002 for Tb and Lu 1995 67

Seawater Mg(OH)2, Al2(OH)3

coprecipitation SPQ-8000 500 10-20 0.13 for Tm 1996 68

River water SPE with chelating resin SPQ-8000A 500 50 0.003 for Tm 1998 69

Stream water Cation-exchange Elan 6000 NAc) NAc) 1 2003 71

River water SPE with chelating resin HP 4500 10 6.7 0.002 for Lu 2003 49

River water LC preconcentration Elan 6000 2 2 0.2 for Tm 2004 73

Groundwater Direct Element II NAc) NAc) 0.007 for Ho 2006 75

River water, seawater

Direct Thermo X-series NAc) NAc) 0.05 for La 2006 76

Seawater, river water

Solvent extraction Thermo X-series 10 8 0.03 for Tb 2007 77

River water Direct Element II NAc) NAc) 0.01 for Pr, Tb, Ho, Er, Tm 2009 78

Rainwater SPE with Chelating resin HP 4500 300 30 0.0017 for Ho 2011 80

Lake water, synthetic seawater

Mg(OH)2 coprecipitation/LA d) Agilent 7500a 40 8-88 0.03 for La, Ce, Pr, Tb 2011 81

Seawater Direct with HMI e) Agilent 8800 NAc) NAc) 1.94 for Dy 2016 85


Lake water, river water, seawater

Magnetic SPE with Fe3O4@TiO2@P2O4

nanoparticles f) Agilent 7500a 50 100 0.01 for Tm 2017 90

Tap water, river water

Calcium alginate microparticlesg) Agilent 7900 100 100 0.01 for Tb, Ho, Er, Tm 2018 94

River water Fe(OH)3 precipitation Agilent 7500cc 1000 33 2.1 for Ho 2018 95

River water SPE with chelating resin Agilent 7700x 50 40 0.001 for Tb and Yb 2019 96

a), Typical sample volume; b), Typical preconcentration factor; c), Not available; d) Measurement of REEs in Mg(OH)2 precipitates by LA-ICP-MS; e)

Measured directly by ICP-MS with high matrix introduction (HMI) mode; f) Lab-prepared Fe3O4@TiO2@P2O4 used as the solid phase for magnetic SPE

preconcentration of REEs, a compound with “@” symbol indicates the surface modification by another compound following the “@” symbol, e.g. “A@B”,

the surface of A modified by B; g) Calcium alginate microparticles used as the solid phase of cation exchange column, which was applied to

preconcentration of REEs


Table 2 The Gd concentration and GdSN/Gd*SN index for natural land water reported in various part of the world.

Natural land water

Year a) Gd

concentration

/ng L-1

GdSN/Gd*SN b) Reference

Name Collected site Country

Havel River Downstream Germany 2009 492 653 45

Tama River Middlestream Japan 2018 141 116 47

Paranoá Lake Branch Brazil 2017 34.4 67.6 57

Jinzhong River Downstream China 2018 86.7 48.4 56

Sakai River Middlestream Japan 2017 36.6 31.6 47

Wupper River Downstream Germany 1995 32.6 30.0 41

Kanda River Middlestream Japan 2002 35.0 28.8 50

Anhumas Creek Middlestream Brazil 2013 56.0 23.6 46

Bobrza River Downstream Poland 2014 46.0 15.3 53

Rhine River Downstream Germany 2011 16.4 15.1 48

Eindergatloop Upstream Belgium 2006 93.0 8.13 62

Weser River Downstream Germany 2005 20.6 8.06 44

Tempaku River Downstream Japan 2001 14.0 7.62 43

Neuse River Upstream United States 2016 97.4 7.11 61

Sagami River Estuarine Japan 2017 5.60 6.14 47

Honokohau Harbor Well Groundwater United States 2012 30.1 5.20 63

Dommel River Upstream Belgium 2007 62.0 4.78 62

Elbe River Downstream Germany 2005 14.9 3.90 44

Yodo River Middlestream Japan 2002 16.6 3.23 51

Kyjsky Pond Czech 2000 4.33 3.14 60

Thames River Middlestream England 2009 4.40 2.90 45

Danube River Upstream Austria 2005 7.63 2.71 45

Atibaia River Downstream Brazil 2013 7.40 2.65 46

Nida River Downstream Poland 2014 19.0 2.24 53

Ems River Downstream Germany 2005 24.2 2.16 44

Vltava river Downstream Czech 2000 6.45 2.13 60

Riacho Fundo Creek Middlestream Brazil 2017 9.54 2.09 57

Jaguari River Downstream Brazil 2013 5.00 2.06 46

Ara River Downstream Japan 1996 4.12 2.04 58

Shonai River Downstream Japan 2001 8.90 1.92 43

Adige River Downstream Italy 2003 4.30 1.89 42

Rokytka Creek Upstream Czech 2000 5.60 1.85 60

Jeju Island Groundwater Korea 2009 68.0 1.59 52

Ibi River Downstream Japan 2001 3.53 1.38 43

Nagara River Downstream Japan 2001 3.87 1.38 43

Tone River Downstream Japan 1996 1.62 1.37 58

Uneticky Creek Upstream Czech 2000 2.87 1.28 60

Wiembach Creek Downstream Germany 2012 34.8 1.25 48

Kiso River Downstream Japan 2001 5.29 1.24 43

Owence Creek Middlestream Australia 2005 10.7 1.18 54

Pioneer River Middlestream Australia 2005 2.89 1.18 54

Hind Well Groundwater United States 2012 6.47 1.13 63

Ueda River Upstream Japan 2001 10.8 1.12 43

a) The year when the samples were collected.

b) The values of GdSN/Gd*SN were estimated form the previously reported analytical data. GdSN is the shale (PAAS)-

normalized value of Gd observed in natural land water, Gd*SN is approximated from the straight line obtained from PAAS-

normalized concentrations of Sm, Tb, and Dy using the least square method.


Table 3 The Gd concentration and GdSN/Gd*SN index for seawater reported.

Seawater

Year a)

Gd

concentration

/ ng L-1

GdSN/Gd*SN b) Reference

Name Collected site Country

Tokyo Bay Coastal Japan 2008 9.10 5.93 49

Osaka Bay Coastal Japan 2002 5.99 4.38 51

San Francisco Bay Coastal United States 2013 26.9 2.87 55

Bahia Coast Coastal Brazil 2016 1.04 1.84 64

North Sea Coastal Germany 2005 2.04 1.79 44

Jade Bay Coastal Germany 2005 2.81 1.63 44

North Pacific Open 2012 6.78 1.59 63

Ise Bay Coastal Japan 2001 2.20 1.53 43

Bangdu Bay Offshore Korea 2009 3.14 1.25 52

Hwasun Bay Offshore Korea 2009 3.30 1.21 52

Antarctic Ocean Open 1995 1.07 1.14 43

Sagami Bay Offshore Japan 1993 1.00 1.14 59

East-China Sea Open 1998 0.38 1.05 15

Japan Sea Coastal Japan 1998 1.90 1.03 79

a) The year when the samples were collected.

b) The values of GdSN/Gd*SN were estimated form the previously reported analytical data. GdSN is the shale (PAAS)-

normalized value of Gd observed in natural land water, Gd*SN is approximated from the straight line obtained from

PAAS-normalized concentrations of Sm, Tb, and Dy using the least square method.


Table 4 The concentration of Gd and GdSN/Gd*SN index from the same sampling sites reported more than twice in different years.

Index

Natural land water Coastal seawater

Havel River Rhine River Tama River San Francisco Bay Tokyo Bay

(Downstream) (Downstream) (Middlestream)

1995a) 2009b) 1995a) 2011c) 1995d) 2018e) 1993f) 2013f) 1996d) 2008g)

Gd concentration

(ng L-1) 106 492 8.60 16.4 21.8 141 3.64 26.9 2.64 9.10

GdSN/Gd*SN 122 653 2.67 15.1 1.80 116 1.36 2.87 3.96 5.93

a) Cited from Ref. 41, b) Cited from Ref. 45, c) Cited from Ref. 48, d) Cited from Ref. 58, e) Cited from Ref. 47, f) Cited from Ref. 55, g)

Cited from Ref. 50.


Table S1 Isotopes of interest and their typical polyatomic spectral interferences. (The number in the

parenthesis following a polyatomic spectral interference is the resolution required for separating from

the isotope of interest.)

Isotope of interest Typical polyatomic spectral interferences 151Eu 134Ba17O (7828), 135Ba16O (9294) 153Eu 135Ba18O (9319), 136Ba17O (8714), 137Ba16O (7456) 157Gd 139La18O (8506), 140Ce17O (8089), 141Pr16O (7333) 159Tb 141Pr18O (8577), 142Ce17O (9366), 142Nd17O (8596), 143Nd16O (7708) 161Dy 143Nd18O (8961), 144Nd17O (9087), 145Nd16O (8274), 144Sm17O (10186) 163Dy 145Nd18O (9585), 146Nd17O (9887), 147Sm16O (8612) 165Ho 148Nd17O (11534), 147Sm18O (10144), 148Sm17O (10076), 149Sm16O (9048) 166Er 148Nd18O (11645), 150Nd16O (11444), 148Sm18O (10168), 149Sm17O (11861),

150Sm16O (9163) 167Er 149Nd18O (10626), 149Sm18O (10626), 150Sm17O (10674), 151Eu16O (9655), 150Nd17O

(13877) 169Tm 152Sm17O (10999), 151Eu18O (11107), 153Eu16O (9344), 152Gd17O (11042) 172Yb 154Sm18O (11455), 154Gd18O (10516), 155Gd17O (11753), 156Gd16O (8886), 156Dy16O

(10003) 174Yb 156Gd18O (9895), 157Gd17O (11030), 158Gd16O (8763), 156Dy18O (9895), 158Dy16O

(8763) 175Lu 157Gd18O (9907), 158Gd17O (9969), 159Tb16O (8522), 158Dy17O (10142)


Figure Captions

Fig. 1 Shale (PAAS)-normalized REE patterns for river waters.

A: REE patterns for pristine river waters without any anthropogenic anomalies of

REEs.

B: REE patterns for German river waters with positive anthropogenic Gd anomaly.

C: REE patterns for Rhine river waters with positive anthropogenic La and Sm in

addition to Gd.

A and B were cited from Ref. 48. C was cited from. Ref.109.

Fig. 2 Structures of the main Gd compounds used as the contrast agent of MRI .

Cited from Ref. 114.

Fig. 3 Chromatogram for Gd compounds in the tap water of Berlin in Germany measured by

ZIC-cHILIC-ICP-MS. Cited from Ref. 120.

Fig. 4 Shale (PAAS)-normalized REE patterns for the shell of freshwater mussel Corbicula

fluminea. Cited from Ref. 123.

A : REE patterns for the shells collected in the Wester River, where there is no industrial

point source, and, in the Rhine River upstream of Worms, where is situated in the

industrial point source.

B : REE patterns for the shells collected in downstream of Worms.

Fig. 5 The average per capita (polulation-weighted) fluxes of 62 elements discharged by

wastewater from WWTF in Switzerlamd. The range of the element fluxes is illustrated

by the color scaling of the periodic table. Cited from Ref. 141.


Fig. 1

C


Fig. 2


Fig.3


Fig. 4


Fig. 5


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