# Rare earth element abundances and gadolinium contamination in tap water worldwide

Jean-Alix Barrat, Germain Bayon, Raphaël Tripier, Marie-Laure Rouget,

Yoan Germain, Douraied Ben Salem

► To cite this version:

Jean-Alix Barrat, Germain Bayon, Raphaël Tripier, Marie-Laure Rouget, Yoan Germain, et al. Rare earth element abundances and gadolinium contamination in tap water worldwide. Chemosphere, 2026, 394, pp.144812. 10.1016/j.chemosphere.2025.144812. hal-05477022

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# Rare earth element abundances and gadolinium contamination in tap water worldwide

Jean-Alix Barrat<sup>a,b,*</sup> <sup>ID</sup>, Germain Bayon<sup>c</sup>, Raphaël Tripier<sup>d</sup>, Marie-Laure Rouget<sup>e</sup>, Yoan Germain<sup>c</sup>, Douraied Ben Salem<sup>f,g ID</sup>

## HIGHLIGHTS

* Determination of the abundance of macrocyclic Gd-based contrast agents in water.

* Gd-based contrast agents can be degraded during tap water production.

* Tap water in large cities is commonly contaminated with Gd.

* The highest levels of Gd contamination are currently observed in Europe.

## GRAPHICAL ABSTRACT

<figure><img src="image_4.png" /></figure>

## ARTICLE INFO

**Keywords:**

Tap water

Gadolinium

Rare earth elements

GBCA

Macrocyclic GBCA

MRI

## ABSTRACT

Gadolinium-based contrast agents (GBCAs) are widely used in magnetic resonance imaging (MRI) and have emerged as persistent contaminants in surface and drinking water since the 1990s. We present a method based on the coprecipitation of rare earth elements (REEs) with iron hydroxides to estimate the amount of macrocyclic GBCAs in water, and apply it here for the first time to investigate the degree of Gd contamination in tap water worldwide. Our results indicate that Gd contamination is widespread in drinking water distribution systems. Tap water in big European cities such as Berlin, Amsterdam, Paris, and Lyon is more contaminated in Gd than tap

<footnote><sup>a</sup> Univ Brest, CNRS, Ifremer, IRD, LEMAR, Institut Universitaire Européen de la Mer (IUEM), rue Dumont d'Urville, Plouzané, F-29280, France</footnote>

<footnote><sup>b</sup> Institut Universitaire de France, Paris, France</footnote>

<footnote><sup>c</sup> Univ Brest, CNRS, Ifremer, Geo-Ocean, Plouzané, F-29280, France</footnote>

<footnote><sup>d</sup> Univ Brest, UMR, CNRS 6521, CEMCA, 6 Avenue Victor Le Gorgeu, 29200, Brest, France</footnote>

<footnote><sup>e</sup> Univ Brest, CNRS, UMS 3113, Institut Universitaire Européen de la Mer (IUEM), F-29280, Plouzané, France</footnote>

<footnote><sup>f</sup> Univ Brest, INSERM, UMR 1101, LaTIM, 22, Avenue C. Desmoulins, Cedex 3, 29238, Brest, France</footnote>

<footnote><sup>g</sup> Brest University Hospital, Hôpital de la Cavale Blanche, Service d'imagerie médicale, Cedex, 29609, Brest, France</footnote>

<footnote>* Corresponding author. Univ Brest, CNRS, Ifremer, IRD, LEMAR, Institut Universitaire Européen de la Mer (IUEM), rue Dumont d'Urville, Plouzané, F-29280, France</footnote>

<footnote>E-mail address: barrat@univ-brest.fr (J.-A. Barrat).</footnote>
water from Asian (Beijing, Guangzhou, Wuhan, Shanghai, Seoul, Tokyo) or American megacities (Los Angeles, New York, and Rio de Janeiro). Furthermore, our data suggest that drinking water treatment processes degrade Gd complexes to varying extent, which can increase the bioavailability of free Gd ions.

# 1. Introduction

From the late 1980s, Gd-based contrast agents (GBCAs) have been routinely used in magnetic resonance imaging (MRI). More than a hundred million MRI examinations are performed each year worldwide, and this number is growing steadily. During magnetic resonance examination, about 800 mg of Gd (for a ~70-kg patient) are typically injected. This dose of gadolinium is then rapidly eliminated, primarily through urine, and subsequently discharged into the sewer system. Because GBCAs are not efficiently treated by wastewater treatment plants, these products are therefore delivered to surface freshwaters, such as rivers, ultimately ending up in coastal marine waters. Since Bau and Dulski (1996) first discovered anthropogenic Gd pollution in German rivers, Gd-contaminated surface water has been measured in various populated areas (e.g., Nozaki et al., 2000; Elbaz-Poulichet et al., 2002; Bau et al., 2006; Liu et al., 2022) and in coastal areas (e.g., Hatje et al., 2016; Pedreira et al., 2018; Ma et al., 2019; Kraemer et al., 2024; Zocher et al., 2025). This pollution has also been found in marine mollusks (Le Goff et al., 2019; Barrat et al., 2022a). The impact of long-term exposure to low amounts of Gd on human health (Souza et al., 2021) and aquatic fauna is still poorly understood, and is a matter of growing interest (e.g., Hanana et al., 2017).

Since the first GBCAs were marketed in 1987, different types of molecules were developed. Linear GBCAs were the first developed, but they were responsible of a potentially fatal pathology (nephrogenic systemic fibrosis) for patients with renal insufficiency (Grobner, 2005; Marckmann, 2006), and possibly to significant Gd accumulation in brain (e.g., Gulani et al., 2017). Consequently, these molecules were gradually phased out after that and officially banned by EU in 2017 (European Medicines Agency). Two linear GBCAs are still occasionally used for very specific examinations, in the absence of any other alternative. In EU, medical examinations are now carried out nearly exclusively with highly stable macrocyclic contrast agents in order to limit possible accumulation in organs.

In Earth's surface environments, dissolved Gd in freshwater is initially geogenic (i.e. derived from natural geological sources), meaning that it is naturally present in the environment. However, this water may also contain substantial anthropogenic Gd linked to medical imaging or any other sources of pollution. Speciation analysis of the latter is likely to exhibit a cocktail of GBCAs, possibly molecules linked to the degradation of these products, and/or Gd directly released from GBCAs. Various protocols have been described to quantify the proportions of the different Gd species present in the environment (Künnemeyer et al., 2009; Raju et al., 2010; Birka et al., 2013; Horstmann et al., 2021; Okabayashi et al., 2021; Athmer et al., 2025; Dalla Costa et al., 2025). However, at present, only few laboratories perform these measurements. In environmental or water quality studies, the proportion of anthropogenic Gd can be easily estimated using REE patterns (e.g., Bau and Dulski, 1996). To date, though, no simple method has been proposed for determining the proportions of chelated and unchelated Gd. Here, we describe an analytical procedure based on ICP-MS that can easily estimate these proportions. We applied this method to a series of tap water samples collected from various cities worldwide. Our goal was to obtain a snapshot of the level of anthropogenic Gd contamination in tap water distributed in selected cities and megacities today, and to evaluate the speciation of this element.

# 2. Samples and analytical procedures

## 2.1. Samples

Seventy-three tap water samples were obtained for this study, collected from sixty large cities and megacities mainly from mid-May to September 2025: thirty-two from Europe, twelve from Asia, seven from North America, two from South America, four from Oceania, two from Africa and one from Lesser Antilles.

Cold tap water sampling was conducted a few minutes after turning on the tap. Samples were collected in acid-cleaned 250 ml polyethylene bottles or in 50 ml polypropylene centrifuge tubes. Clean tubes or bottles were rinsed with tap water several times before samples were taken. No acid was added and no filtration was performed, so that the samples correspond exactly to the water delivered by the tap. The samples were stored in cold conditions, in a fridge whenever possible, and analyzed a few days after collection.

## 2.2. Determination of REE and Y concentrations in water

The procedure used here is directly derived from the method of Bayon et al. (2011), initially developed for the analysis of river and seawater, but recently adapted for industrial beverages (Bordeneuve-Pérès et al., 2025). This method is based on the co-precipitation of rare earths and Y (REY) with Fe hydroxides, after addition of a Tm spike (Barrat et al., 1996). For each water sample, two aliquots (each about 45–50 g) are weighed precisely and spiked with a Tm solution. The first one (aliquot A) is completely evaporated in a Teflon beaker on a hot plate at 110 °C. The residue is taken up with approximately 5–7 drops of 14 N HNO₃. The beaker is then closed and left on hotplate at the same temperature for a few hours to oxidize and decompose the chelating ligands, thereby releasing Gd³⁺ from any contrast agents present. After evaporation on hotplate, the residue is then taken up in 15 ml of ultrapure water. The second aliquot (aliquot B) is left on hotplate for a few hours to reduce its volume to 15–20 ml to facilitate handling. Approximately 300 µg of Fe (0.3 ml of a PlasmaCAL Fe solution at 1000 ppm) is added to each aliquot. After homogenization, 0.75 ml of concentrated ammonia solution is added to co-precipitate Fe-hydroxides, resulting in efficient scavenging of reactive elements such as REY (Bayon et al., 2011). The beakers are closed and placed on hotplate at 140 °C for 4 h. After cooling, the solutions are centrifuged, and the supernatant is removed and replaced with ultrapure water. The operation is repeated four times. The Fe-hydroxides are then transferred to a beaker, dried, and taken up with 2 ml of ultrapure 0.4 N HNO₃. The REY concentrations are determined with this solution by ICP-MS at the Pôle Spectrométrie Océan (PSO), Plouzané, using a Thermo Scientific sector-field ELEMENT XR™ mass spectrometer. Each solution is analyzed in triplicate, using the same analytical settings and interference corrections as Charles et al. (2021).

Bayon et al. (2011) already demonstrated the reliability of this method using a series of natural water reference materials. Here, we used an in-house standard (water “OBH#1”) produced using 1 L of ultra-pure MilliQ water to which we added the equivalent of approximately 3.5 mg of BHVO-2, a basalt from Hawaii, contained in a few drops of 6 N HCl. We chose BHVO-2 because it is one of the most widely used geostandards and its concentrations for major and trace elements are accurately known (e.g., Jochum et al., 2016). We determined the concentrations of numerous elements (e.g., Fe, Mg, Sr, Mn, Ca) in water OBH#1 by ICP-MS, which enabled us to determine precisely the dilution factor of BHVO-2 in OBH#1 and, in turn, to deduce its theoretical REY
<figure><img src="image_5.png" /><figcaption>Fig. 1. Proportions of chelated Gd contained in water, coprecipitated with Fe hydroxides after addition of ammonia. See text for details.</figcaption></figure>

concentrations. We analyzed OBH#1 ten times using the procedure above (of course, like an aliquot “B” because OBH#1 is devoid of GBCA), and the abundances we obtained are in excellent agreement with theoretical values (Table S1). The relative standard deviations are better than 5 % for all elements.

In this work, the data will be represented using rare earth + Y patterns, normalized to the Post-Archean Australian Shale (PAAS), which allows, among other things, the definition of anomalies for certain elements. Readers who are not familiar with this type of diagrams are referred to review papers summarizing their properties and their utility in environmental studies (e.g., Barrat and Bayon, 2024). We use the PAAS values obtained by Pourmand et al. (2012), adjusted to standard results obtained in our laboratory for elemental abundance normalization (Barrat et al., 2020). The La, Ce, and Gd anomalies are calculated geometrically using the La/La*, Ce/Ce*, Gd/Gd* ratios, where X* is the extrapolated concentrations for a smooth PAAS-normalized REE pattern and Xsn is the concentration of element X normalized to PAAS: LAsn* = Prsn^3/Ndsn^2, CEsn* = Prsn^2/Ndsn, (e.g., Barrat et al., 2023).

In the case of freshwaters, Kulaksiz and Bau (2007) showed that Gd behaves like a light REE, rather than a heavy REE. Consequently, and in agreement with these authors, we extrapolate Gd* from the abundances of Sm and Nd (Gdsn* = Smsn^2/Ndsn). Other possible interpolations or extrapolations give fairly similar values for Gd* (not shown). The method used to calculate Gd* does not affect our conclusions because, in most cases, positive anomalies in Gd are quite clear or even very significant.

Based on the OBH#1 standard (Table S1) and other reference materials, the inferred reproducibility (RSD) for concentrations and elemental ratios (including Gd/Gd*) is generally better than 5 %.

## 2.3. Is Gd in GBCA co-precipitated with Fe-hydroxides?

The method developed by Bayon et al. (2011) was initially designed for the analysis of GBCA-free natural waters, so a pre-requisite was first to evaluate the behavior of GBCAs during Fe-hydroxide coprecipitation.

Today, four macrocyclic GBCAs (gadoteric acid, gadoteridol, gadobutrol, and gadopiclenol) and two linear GBCAs (gadobenic acid, gadoxetic acid) are used by European medical imaging services. These products were provided to us by the Brest University Hospital in the form of commercial injectable solutions. To evaluate their behavior, we used our internal OBH#1 standard, which we doped with GBCAs to obtain a strong positive anomaly in Gd (Gd/Gd* of around 100). We performed at least two experiments per GBCA: we first analyzed each solution according to the protocol described above (section 2.1) with Fe hydroxide co-precipitation at 140 °C for 4 h, and a variant of this protocol with Fe co-precipitation taking place at room temperature for seven days. The results are shown in Tables S2-S3 and Figs. S1-S3. The Gd concentrations in aliquots A and B (GdA and GdB) for each experiment allow the proportion (P) of Gd contained in the contrast agents that are co-precipitated with Fe hydroxides to be calculated directly:

$$P = (Gd_B - Gd_{OBH#1}) / (Gd_A - Gd_{OBH#1}) \quad (1)$$

Both linear GBCAs are almost completely co-precipitated with Fe hydroxides, while the four macrocyclic GBCAs are only slightly co-precipitated (Fig. 1). This behavior is not temperature-dependent, as both room temperature and 140 °C experiments yielded similar results. We verified the result obtained for gadobutrol, i.e. the macrocyclic GBCA that co-precipitates most efficiently with Fe hydroxides (≈5 %), by doping our OBH#1 water with a different batch of this contrast agent and increasing the doping by an order of magnitude (Gd/Gd* in the order of 1000). These experiments confirm that the behavior of this GBCA is the same for both batches, and not dependent on solution concentrations. One possible explanation to account for the different behaviors of macrocyclic and linear GBCAs during Fe-hydroxide co-precipitation would relate to transmetalation processes. Linear GBCAs are indeed prone to transmetalation by Fe³⁺ in contrast to macrocyclic GBCAs (Telgmann et al., 2012; Rabiet et al., 2014). This would require further experimental work, beyond the scope of the present study.

These observations have important implications. First, the method for determining REYs by coprecipitation with Fe hydroxides cannot be applied directly to water containing GBCAs. To correctly determine Gd abundances, it is essential to degrade the contrast agents beforehand,
<figure><img src="image_6.png" /><figcaption>**Fig. 2.** REY patterns for three tap waters illustrating the method for estimating the quantities of macrocyclic GBCA agents contained. Since linear GBCAs are virtually no longer used in EU, anthropogenic Gd distinct from macrocyclic GBCAs (shown in yellow in this figure) necessarily derives from the degradation of the latter.</figcaption></figure>

which is done with HNO₃ for our aliquot A. This behavior of GBCAs could be seen as a weakness of the coprecipitation method. On the contrary, the limited co-precipitation of macrocyclic GBCAs with Fe hydroxides (<< 5 % for most of them) allows their presence in water to be easily detected and even their abundance to be reasonably well estimated (see below).

## 2.4. Geogenic and anthropogenic Gd in water

Contamination of water by anthropogenic Gd is indicated by the presence of a positive anomaly in Gd, which is clearly visible in its REE pattern. It can be quantified by calculating Gd anomalies, where Gd/Gd* significantly >1 points toward anthropogenic contamination. There is no agreement among authors on the maximum value that the Gd/Gd*
<figure><img src="image_7.png" /><figcaption>Fig. 3. Selected REY patterns of tap water normalized to PAAS.</figcaption></figure>

ratio can yield in uncontaminated water and materials. Gd/Gd* ratios slightly higher than 1.35 have been measured, for example, in present-day biogenic carbonates free of any contamination (Barrat et al., 2022a,b). Despite these uncertainties, a good estimate of the amount of geogenic (i.e., natural) Gd is given by Gd*. Consequently, the amount of anthropogenic Gd (Gd<sub>anthr</sub>) contained in a sample is simply given by the following relationship (e.g., Bau and Dulski, 1996):

$$Gd_{\text{anthr}} \approx Gd_{\text{total}} - Gd^* \quad (2)$$

Even though there may be some uncertainty in estimating the amount of geogenic Gd, this is not particularly significant. If the sample is slightly contaminated or not contaminated at all, the Gd<sub>anthr</sub> value will be low, and this will not bias the interpretation in any case. If, on the other hand, the contamination is more significant, Gd* will have no impact on the conclusions.

Medical imaging departments may use different contrast agents. Consequently, natural waters contaminated with anthropogenic Gd may contain a cocktail of GBCAs. In theory, this water may contain macrocyclic GBCAs (Gd<sub>m-GBCA</sub>), linear GBCAs (Gd<sub>l-GBCA</sub>), and finally Gd resulting from the degradation of these two types of contrast agents or eventually anthropogenic Gd from other sources (Gd<sub>d-GBCA</sub>):

$$Gd_{\text{anthr}} = Gd_{\text{m-GBCA}} + Gd_{\text{l-GBCA}} + Gd_{\text{d-GBCA}} \quad (3)$$

Concentrations obtained with our two aliquots A and B allow us to estimate the quantities of these different Gd compounds contained in water (Fig. 2). The Gd concentration measured with aliquot A after treatment with nitric acid corresponds to the total Gd concentration contained in the water (Gd<sub>total</sub>). The experiments we conducted with the different GBCAs indicate that the Gd concentration measured with aliquot B (which we will denote Gd<sub>coFe</sub> for clarity, with “coFe” for coprecipitated with Fe hydroxides) is close to the abundance of Gd not carried by macrocyclic GBCAs. Consequently:

$$Gd_{\text{m-GBCA}} \approx Gd_{\text{total}} - Gd_{\text{coFe}} \quad (4)$$

and

$$Gd_{\text{coFe}} - Gd^* \approx Gd_{\text{l-GBCA}} + Gd_{\text{d-GBCA}} \quad (5)$$

In the event of very slightly negative values of Gd<sub>total</sub> - Gd<sub>coFe</sub> and Gd<sub>coFe</sub> - Gd* (within the limits of analytical uncertainties), these are not considered significant and are assumed to be equal to 0.

In EU, consumption of linear GBCAs has declined drastically since 2005 (Le Goff et al., 2019). Most of these molecules have been banned since 2017, and only the two linear GBCAs tested in our experiments are still being administered. These two products are used occasionally to detect certain pathologies, and their presence in the environment is consequently negligible. We can therefore assume that today in EU, Gd<sub>l-GBCA</sub> ≈ 0. This assumption may also hold to North America but this would require confirmation. We are not aware of any recent public statistics on the consumption of the various contrast agents at the national level outside EU or at the global level.

Based on the results for OBH#1 and the reproducibility of the runs obtained on the different water samples, the reproducibility (RSD) for (Gd<sub>coFe</sub> - Gd*), Gd<sub>m-GBCA</sub>, and Gd<sub>anthr</sub> is much better than 10 %.

### 3. Results

The sample list and abundances of REY in tap water are given respectively in Tables S4 and S5. These elements are of low solubility in water (e.g., Sholkovitz, 1995), and consequently it is not surprising that their concentrations are low in tap water (ΣREE = 2–79 ng/kg). All the studied samples contained traces of Ba, which generated isobaric interference from Ba hydroxides on Eu. This interference was too significant to be corrected satisfactorily. For this reason, we have not reported the Eu concentrations measured in tap water.
<figure><img src="image_8.png" /><figcaption>Fig. 4. Macro cyclic GBCAs vs. anthropogenic Gd in tap water.</figcaption></figure>

The REY patterns for water samples from different regions world-
wide are reported in the supplementary materials (Figs. S4-S25). Only a
selection is shown in Fig. 3. The observed REY patterns vary in shape.
Most REY patterns are heavy REE enriched ((Tb/Yb)sn = 0.01–1). They
show variable anomalies in Y ((Y/Ho)sn = 0.4–3.1), and negative
anomalies in Ce for all but one samples (Ce/Ce* = 0.08–0.9). This latter
exception is Rio de Janeiro, which shows a large positive anomaly in Ce
(Ce/Ce* = 3.0), the anthropogenic origin of which is beyond any doubt
but yet to be determined. The REE patterns of tap waters show no large
positive anomalies in La (La/La* = 0.7–3.5) or Lu (Lu/Lu* = 0.4–1.4)
that could be attributed to an anthropogenic signature (Kulaksiz and
Bau, M., 2011b; Aktar and Toyoda, 2024; Alemu et al., 2025). However,
almost all samples show marked positive anomalies in Gd (Gd/Gd* =
0.6–125), clearly reflecting substantial Gd contamination from medical
imaging. Only a few samples have Gd/Gd* ratios <1.2, suggesting that
they are not contaminated or are only minimally contaminated. This
includes tap waters from Stanford (USA, California), which originates
from the Tuolumne River watershed, Yosemite National Park), from Kiel
(Germany), Beijing (China), New York City (USA), Wellington (New
Zealand), Pointe-à-Pitre (Lesser Antilles), and Merlo (Argentina).
Moreover, traces of macrocyclic GBCAs are detectable, particularly in
the most contaminated samples (Fig. 4).

4. Discussion

4.1. *Origin of the shapes of REY patterns*

The water used for industrial tap water production comes from various sources, such as river water, groundwater, or even seawater. These natural waters have different initial REY concentrations, which explains part of the diversity of the patterns. For some cities, the water supplied may come from different sources depending on the neighborhood. This is clearly illustrated by Paris and its suburbs, where water is mainly produced from the Seine and two of its tributaries (the Oise and the Marne), supplemented by groundwater and water from more distant river sources (the Voulzie, Vanne, Loing, and Avre River) which are brought to Paris by aqueducts. This diversity of supply partly explains the variability in REY abundances in tap water from the Paris Region (Fig. 3a). Berlin and Amsterdam are also particularly illustrative examples. For historical reasons, the organization of the water distribution network differed between East Berlin and West Berlin before reunification. Even today, the waters of West Berlin are still characterized by much greater Gd anomalies than those of the eastern sector of the city (Kulaksiz, and Bau, 2011a; Tepe et al., 2014; Krohn et al., 2024). For Amsterdam, there are two sources of water used to produce tap water. The part of the city located west of the Amstel River is supplied with water produced from the Rhine, while in the east, the water comes from a polder located north of Utrecht. These waters do not have the same REY abundances.

In addition to this diversity of sources, the purification processes can
significantly alter the concentrations of trace elements in water. Indus-
trial water purification is carried out through a series of steps that
include filtration with sand or activated carbon, clarification, decanta-
tion, disinfection (particularly through ozonation, addition of chlorine,
and UV exposure), and sometimes reverse osmosis (e.g., Lahnsteiner,
2024). These steps may be accompanied by the addition of FeCl3 (for-
mation of FeO(OH) flocs to remove suspended particles) and caustic
soda to reduce limestone concentrations. In addition, limewater (solu-
tion of Ca(OH)2) may also be added for pH correction, to remove traces
of metals, or even to mineralize the water to be distributed. To the best
of our knowledge, the effects of these different steps on REY concen-
trations in water are not described in the literature.

However, some of our samples allow us to *qualitatively* assess the effect of the entire purification process on REY abundances in tap water. Certain cities, such as Seoul, Nantes, and Budapest, are supplied by a single water source, namely the Han, Loire, and Danube rivers, respectively, for which data are available (Song et al., 2017; Rétif et al., 2024; Zhang et al., 2024). Amsterdam was added to these three cities because two of our samples comes from a neighborhood west of the Amstel River, which is supplied with water produced from the Rhine River. The composition of this river was determined in its Dutch section (Klaver et al., 2014), and also in Germany north of Bonn (Kulaksiz and Bau, 2007). Tap water was compared to river water in Fig. 5. Several important observations emerge:

- The REY pattern of tap water is much more fractionated than the corresponding initial river water. In all four cases, the heavy REE concentrations in tap water are comparable to those in river water, but light REE concentrations are much lower. Similar observations were made for tap water in Osaka (Inoue et al., 2025) and Berlin (Kulaksiz and Bau, 2011). The formation of Fe hydroxide (either as floc during a water clarification step with Fe chloride as suggested by Kulaksiz and Bau (2011a), or as precipitates during an ozonation step), could effectively co-precipitate light rare earth elements preferentially.

- In Nantes, Seoul, or Budapest, tap water has much more pro-
nounced negative Ce anomalies than each corresponding river water.
Water treatment involves usually some highly oxidizing steps (ozona-
tion, addition of Cl), to decompose complex organic compounds
(including pollutants), to remove certain metals (e.g., Fe²⁺/³⁺, Mn²⁺), as
well as algae, viruses, and bacteria. During these steps, part of the Ce³⁺
<figure><img src="image_9.png" /><figcaption>Fig. 5. Comparison between tap water and river water used for production (Danube River from Zhang et al. (2024), Han River from Song et al. (2017), Loire River from Rétif et al. (2024), and Rhine from Klaver et al. (2014)).</figcaption></figure>

in solution can be oxidized to insoluble Ce⁴⁺ (e.g., Bauer and Lindstrom, 1968), explaining the decoupling of Ce from other REEs and the development of marked negative Ce anomalies. Surprisingly, the Rhine and tap water in Amsterdam have similar negative Ce anomalies (Ce/Ce* ≈ 0.7). We do not explain this singularity, since this water has been ozonated, but no chlorine was added at the end of the purification process. More data would be necessary to discuss this feature.

- In northern Germany and in Netherlands, the Rhine River water displays a characteristic positive La anomaly, linked to industrial pollution from a production plant for fluid catalytic cracking catalysts used in petroleum refining (Kulaksiz and Bau, 2011b; Klaver et al., 2014). This anomaly is no longer detectable in tap water, suggesting that the purification process has effectively removed the fraction of anthropogenic La.

- The river water samples used to compare tap water were collected some years before the tap water samples (e.g., 2010 for Rhine water, Klaver et al., 2014). As Gd levels in rivers reflect the growth in GBCA consumption (e.g., Lerat-Hardy et al., 2019; Inoue et al., 2020), the Gd concentrations reported for rivers in Fig. 5 are hence only indicative. It should nevertheless be noted that the concentrations of Gd in these four tap waters are lower, although of the same order of magnitude, than those found in river water. For these four examples, water purification does not effectively remove GBCAs, but the data do not allow us to rule out the possibility that a fraction of them may have been removed.

## 4.2. Degradation of GBCAs during the production of tap water

The studies conducted on Gd speciation in wastewater treatment plant effluents have shown that GBCAs are not degraded by treatment and are therefore discharged into rivers (e.g., Dalla Costa et al., 2025; Athmer et al., 2025). The speciation of Gd in contaminated water used to produce distribution water does not show any significant amounts of Gd

resulting from the degradation of GBCAs. The question of the possible degradation of GBCAs during the production of tap water must be evaluated.

The technique we have developed enables us to estimate the amount of macrocyclic GBCAs contained in water and compare it with the amount of total anthropogenic Gd (Table S5, Figs. 4 and 6).

Macrocyclic GBCAs alone do not explain the positive Gd anomaly in tap water. First, some samples exhibiting high Gd/Gd* ratios indicative of significant anthropogenic contamination may be free of them: as for example, tap water from Toulouse (Gd/Gd* = 4.9, Fig. 2c), Brussels (Gd/Gd* = 2.8), Belem in Lisbon (Gd/Gd* = 1.8), Los Angeles (Gd/Gd* = 3.9–8.2), Guangzhou (Gd/Gd* = 4.5), Wuhan (Gd/Gd* = 7.9), and Shanghai (Gd/Gd* = 4.8). For the other samples in Europe or elsewhere, macrocyclic GBCAs account for 11–95 % of anthropogenic Gd (Fig. 4b).

These observations raise the question of the speciation of the non-macrocyclic fraction of anthropogenic Gd contained in tap water. Since most linear GBCAs have been banned in EU, they are no longer released into the environment. Indeed, they are no longer detected in surface waters in Germany (Macke et al., 2021; Athmer et al., 2025), whereas they were still present in high concentrations there in 2014 (Birka et al., 2016a).

The proportion of linear GBCAs certainly cannot be invoked to explain the non-macrocyclic fraction of anthropogenic Gd in tap water, at least in EU and United Kingdom (the ban on linear GBCAs was imposed before Brexit, ca. 2020). Birka et al. (2016a) quantified the abundances of GBCAs in six tap waters in the Ruhr metropolitan area. A significant difference between total Gd and GBCAs was obtained, with no evidence of transformation products being present in the samples analyzed. Their data suggest that 18–34 % of anthropogenic Gd present in water was dechelated during its purification. Our results are consistent with this and suggest that macrocyclic GBCAs are partly degraded during water purification, releasing a significant proportion of Gd³⁺ into
<figure><img src="image_10.png" /><figcaption>Fig. 6. Anthropogenic Gd for the tap water samples analyzed in this work.</figcaption></figure>

solution. Oxidative processes such as ozonation and chlorination may be responsible for this degradation. The effect of disinfection by UV irradiation could also be considered (Birka et al., 2016b; Alvarez-Aguiñaga et al., 2022). Experimental studies, as well as analyses of water samples taken from treatment plants throughout the process, are now needed to identify the treatment steps responsible for the dechelation of GBCAs, and to model the concentrations of REYs during the purification of water.

### 4.3. Anthropogenic Gd in tap water. A global or regional problem?

Nowadays, surface waters are frequently contaminated with Gd in populated areas where access to modern medical technology is readily available. As these waters are used for industrial production of drinking water, questions about the impact of Gd on human health are regularly raised (e.g., Thomsen, 2017; Brünjes and Hofmann, 2020; Souza et al., 2021). To discuss this topic, it is necessary to establish the current state of drinking water contamination on a global scale. Prior to our study, the
data available in the literature were geographically limited. Most of the high-quality data previously published concerned only a few regions in Germany (Kulaksız and Bau, 2011a; Tepe et al., 2014; Schmidt et al., 2019; Krohn et al., 2024), Poland (Wysocka et al., 2023), and Japan (Inoue et al., 2025). Our new set of data provides a much broader picture of anthropogenic Gd contamination in tap water worldwide, albeit with some limitations. First, only two samples from Africa (Alger and Tunis) were analyzed here, which complement one previous study reporting evidence for the occurrence of anthropogenic Gd contamination in urban waters from Benin (Atinkpahoun et al., 2020). However, Africa still remains poorly equipped in terms of MRI facilities (e.g., Obole et al., 2018), so it is unlikely that anthropogenic Gd is found ubiquitously across this continent. Additionally, we obtained only two samples from South America, and no data is available for tap water from India, where Gd contamination is much more probable.

We have represented the integrity of our dataset in Fig. 6. Our samples were all collected over a 4.5-month interval in 2025, which gives us a broad idea of the present-day situation regarding anthropogenic Gd contamination worldwide. The degree of Gd contamination varies considerably from one city to another, but it is strikingly apparent that it is higher in Europe compared to the rest of the world. In European cities, water contamination levels frequently exceed 4 ng/kg (for 23 out of 43 samples), with values even higher than 60 ng/kg being measured in Berlin and Amsterdam. These levels contrast with those observed elsewhere, as only 3 of our 20 samples outside Europe exceed 4 ng/kg (Tunis, Rio de Janeiro, and Washington, D.C.). This observation is significant because our sampling collection features some of the largest cities by population worldwide, including Istanbul (15 million inhabitants) Paris (12 million including adjacent metropolitan areas), London and greater London (9 million), Beijing (22 million), Wuhan (13 million), Guangzhou (15 million), Shanghai (28 million), Tokyo (37 million), Seoul (25 million), New York City (20 million), Rio de Janeiro (6.7 million) and Los Angeles (18 million). We did not include the data from the literature cited above because these water samples were taken a few years or more before our study. However, these previous results are in perfect agreement with the trend shown in Fig. 6. Thus, the differences in tap water quality, at least for Gd, between Europe and the rest of the world are striking and significant.

The difference between Europe and the rest of the world in terms of Gd contamination cannot be explained simply. Explanations must be sought on a case-by-case basis. If we take all the data, contamination in tap water is not correlated with the number of inhabitants in each city, or with the number of MRI facilities (not shown). Contamination in tap water depends primarily on the choice of resources used to produce it, and therefore on the level of exposure of that resource to effluents containing Gd. This is clearly illustrated by certain large European cities such as Amsterdam, Paris, and London, where tap water is mainly produced from river water known to be moderately to highly contaminated with Gd. Conversely and obviously, certain cities such as Stanford and Kiel, which use resources that are little or not at all contaminated, have tap water that is free of anthropogenic Gd.

There is a second explanation relating to water treatment. It can be assumed that differences in water treatment lead to marked differences in the behavior of GBCAs and their degradation. This could also explain the low concentrations of anthropogenic Gd in some Asian cities. For example, the results obtained for the water in Wuhan, Guangzhou, and to a certain extent Shanghai are quite surprising. Not only are anthropogenic Gd amounts low, but macrocyclic GBCAs are either undetectable or in very low quantities. We do not know which GBCAs are injected in China, but it would be surprising if linear GBCAs were still widely used in this country. If this is not the case, the absence or rarity of macrocyclic GBCAs would imply that they have been very effectively degraded during water purification, and why not eliminated during one of the purification stages. This possibility is currently only speculative. We reiterate the need to study the behavior of REYs and GBCAs during water purification.

# 5. Conclusions

In aqueous solution, macrocyclic GBCAs are not or only slightly coprecipitated with Fe hydroxides. This property makes it easy to determine the abundance of these GBCAs in water using ICP-MS, without any special dedicated speciation analysis. The method proposed here can be used directly by all laboratories already measuring REYs in water and should quickly enable the multiplication of this type of data in both surface water and industrial effluents. Our procedure opens important prospects for describing and modeling the behavior of REYs and GBCAs during natural processes, but especially during industrial processes (water treatment).

This method has been applied to tap water samples taken in large cities and megacities in Europe, Asia, and America. Most of the samples in this study show a positive anomaly in Gd, confirming the extent of Gd contamination in the water used to produce tap water. Macro cyclic GBCAs represent only a fraction of anthropogenic Gd in tap water, implying that a significant fraction of them has been degraded during water purification processes.

Finally, our data suggest that Gd contamination of tap water in Europe appears to be higher than elsewhere in Asia or North America. This observation would need to be confirmed by further measurements, particularly on tap water from other large cities and megacities in North and South America and in India, for which little or no data are available. If this trend is established, it would raise critical issues regarding medical practices (level of hospital equipment, use of Gd for medical imaging), wastewater treatment and water resource conservation.

## CRediT authorship contribution statement

**Jean-Alix Barrat:** Writing – review & editing, Writing – original draft, Visualization, Validation, Supervision, Resources, Project administration, Methodology, Investigation, Funding acquisition, Formal analysis, Data curation, Conceptualization.
**Germain Bayon:** Writing – original draft, Visualization, Validation, Methodology, Conceptualization.
**Raphaël Tripier:** Writing – original draft, Validation.
**Marie-Laure Rouget:** Methodology, Investigation.
**Yoan Germain:** Methodology, Investigation.
**Douraied Ben Salem:** Writing – original draft, Validation, Investigation, Conceptualization.

## Declaration of competing interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

## Acknowledgments

We thank Professor M. Horvat for the editorial handling, and the anonymous reviewers for their constructive comments. This work was supported by an IUF grant to JAB, and would not be possible without the help of family, friends and colleagues who collected the samples of tap water (P. Allain, P., E., J. et M. Barrat, M. Bruguet, L. Chauvaud, Y. Choi, J. Collot, E. Cornil, S. Fagot, F. Floc'h, G. Guéguen-Hallouët, A. Guillot, G. Hammouda, G. Hamon, C. Iphar, A. Lopin, W. Lee, C. and F.X. Le Bourdonnec, V. Le Garrec, M. et S. Le Goff, A. Lopin, L. Marié, F.A.C. Le Moigne, P. Nonnote, F. Olivier, A. Pomade, M. Rabineau, P. Rimelin, C. Rivière Paya, R. Ravasse, F. Rioual, C. Saccardy, D. Spenato, A. Tetik, F. Tissot, the family Wallaert, X. Wang, A. Yamaguchi).

# Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.chemosphere.2025.144812.
# Data availability

All the data are in the supplementary materials

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