# Drivers of rare earth elements (REEs) and radionuclides in changing subarctic (Nunavik, Canada) surface waters near a mining project

H. Marginson, G.A. Macmillan, M. Wauthy, E. Sicaud, J. Gérin-Lajoie, J.-P.

Dedieu, M. Amyot

► To cite this version:

H. Marginson, G.A. Macmillan, M. Wauthy, E. Sicaud, J. Gérin-Lajoie, et al. Drivers of rare earth elements (REEs) and radionuclides in changing subarctic (Nunavik, Canada) surface waters near a mining project. Journal of Hazardous Materials, 2024, 471, pp.134418. 10.1016/j.jhazmat.2024.134418. hal-04922970

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# Drivers of rare earth elements (REEs) and radionuclides in changing subarctic (Nunavik, Canada) surface waters near a mining project

H. Marginson<sup>a</sup>, G.A. MacMillan<sup>a</sup>, M. Wauthy<sup>a</sup>, E. Sicaud<sup>b,d</sup>, J. Gérin-Lajoie<sup>c,d</sup>, J.-P. Dedieu<sup>d,e</sup>, M. Amyot<sup>a,d,*</sup>

## HIGHLIGHTS

* First insights into REEs, U and Th in two subarctic river basins before REE mining.

* Modelled field data predicted lower dissolved REEs with higher water temperatures.

* REE distribution and Ce fractionation driven by colloidal Al, Fe and Ti.

* Suspected co-transport of REEs, Th, Al and Fe with humic-rich terrestrial DOC.

* Changes in DOM release to northern systems may alter aquatic REE geochemistry.

## GRAPHICAL ABSTRACT

<figure><img src="image_3.png" /></figure>

## ARTICLE INFO

**Keywords:**

Lanthanides

Arctic

Ce anomaly

Dissolved organic matter (DOM)

Trace metals

Climate change

Thermokarst

Speciation

## ABSTRACT

The emergence of mining projects for rare earth elements (REEs) in response to rising global demand and geopolitical factors introduces environmental concerns, such as the suspected release of anthropogenic REEs to aquatic systems and the coexistence of radionuclides (U, Th). Northern regions confront heightened challenges from limited research and accelerated climate change. Drivers of REEs in surface waters (including George and Koroc rivers, their tributaries, and thermokarst lakes) were studied (2017–2023) in subarctic Canada within a climate transition zone, near a prospective REE mine. Dissolved REEs (<0.45 µm) correlated positively with Al, Fe, Th, U, Cl⁻ and DOC. A novel relationship with water temperature demonstrated an approximate 10-fold decrease in REE concentrations over the environmental gradient (2–20 °C), suggesting complex implications for REE speciation under climate pressures. Optical analyses further predicted REEs were mobilized by humic-rich, terrestrial DOC, with correlations presenting a possible co-transport with Al, Fe and Th. Relationships for redox-sensitive Ce anomalies (Ce/Ce* = 0.18–1.2) with multi-valent trace metals (Al, Fe, Ti) and DOC were

<footnote><sup>a</sup> GRIL, Département de sciences biologiques, Complexe des Sciences, Université de Montréal, Québec, Canada</footnote>

<footnote><sup>b</sup> Département de géographie, Université de Montréal, Canada</footnote>

<footnote><sup>c</sup> Université du Québec à Trois-Rivières, Québec, Canada</footnote>

<footnote><sup>d</sup> Centre d'Etudes Nordiques, Québec, Canada</footnote>

<footnote><sup>e</sup> Institut des Géosciences de l'Environnement (IGE), Université Grenoble-Alpes et CNRS, France</footnote>

<footnote>* Corresponding author at: GRIL, Département de sciences biologiques, Complexe des Sciences, Université de Montréal, 1375 Avenue Thérèse-Lavoie-Roux, Montréal, H2V 0B3 QC, Canada.</footnote>

<footnote>E-mail address: m.amyot@umontreal.ca (M. Amyot).</footnote>
suggestive of a preferential adsorption of Ce by inorganic colloids in low-DOC systems. Findings emphasized the potential for changes in REE geochemistry with ongoing northern surface warming and vegetation shifts.

# 1. Introduction

Rare earth elements (REEs) are critical metals in industrial processes and technological devices, serving as efficient magnets, catalysts, and in metal alloys, among other applications [1]. Over recent decades, the global demand for REEs, which include the 15 lanthanides ($Z = 58–71$) and yttrium ($Z = 39$), has been increasing, and projections suggest this trend will continue [2]. The growing demand has triggered global mining development, driven further by geopolitical pressures, including a market dominated by China, historical pricing instability, and the absence of effective REE replacements or recycling (<1% of REEs) [3,4]. Mining these ‘emerging contaminants’ [5] introduces additional concerns, particularly regarding the release of REEs and co-occurring radionuclides (U, Th) to surface systems [6], where they can bioaccumulate in the food chain and pose toxicological risks [3,7,8].

Canada is among nations with significant REE mineral reserves, featuring one operational mine in Northwest Territories since 2019, and several ongoing projects [9]. As numerous REE deposits are in subarctic and arctic zones of the country, the vulnerability of these regions to multiple environmental pressures is highlighted. Notably, surface warming in the Arctic exceeds the global rate by at least double [10], with anticipated impacts including permafrost degradation, thermokarst formation, and increased tree cover, as observed in the study area of northeastern Canada [11,12]. Consequentially, the release of stored carbon during ground thaw is expected to increase organic matter input to aquatic systems and may potentially remobilize accumulated metals [13]. Further, while the structure of dissolved organic matter (DOM) is complex, it contains humic substances (e.g., humic acid, fulvic acid) that are known ligands of certain trace metals (TMs), including the REEs [14]. In circumpolar North, regions undergoing permafrost thaw have demonstrated increasing fractions of terrestrial-derived (‘allochtho- nous’) chromophoric DOM (c-DOM), studied by Parallel Factor (PAR- AFAC) analysis [15]. This northern shifting of DOM composition has uncertain implications for REE behavior in the environment.

In natural waters, REEs are often attributed to colloids in the dissolved fraction (<0.45 µm), such as with Fe-oxides and organic colloids, due to their high reactivity and low solubility of aqueous species [16-18]. Field studies have presented various influences on REE distribution and speciation, including the environment's pH [19,20] and ionic strength [21,22], precipitation events [17], and regional geology [23]. No consistent trend has been established for the influence of surface water temperature on REEs. Partitioning among individual REEs can arise from an electronic effect that causes decreasing atomic size with increasing atomic number [1]. This contributes to a greater stability for heavy REEs (HREEs, Tb-Lu + Y) in dissolved complexes (e.g., carbonates, DOM phenols) than light REEs (LREEs, La-Gd), which display higher particle reactivity [24,25]. Additionally, while REEs are mostly in the 3+ state, anomalies may naturally develop for elements with multiple stable valences (i.e., Eu, Ce). In particular, Ce anomalies have been associated with aquatic geochemical processes, such as Mn marine redox chemistry [26] and REE adsorption on Fe-oxyhydroxides [27].

The Strange Lake B-Zone deposit in subarctic Canada (Nunavik) is currently in pre-feasibility study and contains one of the world's largest REE reserves, notably enriched in the highly demanded HREEs [28], additionally hosting a co-mineralization of the actinides uranium and thorium [6]. The Strange Lake complex is in the George River Basin (GRB), which makes up the study area with the Koroc River Basin (KRB), though no mining is forecasted in the latter. Through a collaborative project ‘Imalirijiit’ with the Inuit Village of Kangiqsualujjuaq in the GRB [29], multi-annual surface water sampling was conducted to explore the

drivers of REEs, $^{238}$U, and $^{232}$Th in various aquatic ecosystems within a climate transition zone. The influence of water temperature and chemical factors (pH, dissolved oxygen, major anions, trace metals, DOM) on REE distribution and fractionation was studied. We predicted REEs would be related to inorganic colloids of metal oxides (Al, Fe, Mn), driven by surface redox interactions and further explored with cerium anomalies; and to humic-rich DOM due to its strong REE affinity, investigated quantitatively and qualitatively (spectroscopy). A second objective presented baseline values for elements of concern (REEs, U, Th) prior to mining onset in the region.

# 2. Experimental

## 2.1. Study Area

The study area (Fig. 1) consisted of the 42,000 km² GRB and the 4010 km² KRB in Nunavik, Canada [30,31]. Hydrogeomorphometric classification modelling demonstrated two main types of sub-watersheds in the GRB: the north is characterized by higher-elevations and mountain regions, and the south has gentler slopes with higher moisture content and more abundant boreal forest [31]. The KRB features a tundra climate with vegetation dominated by lichens and mosses, and tree growth mostly limited to the river valley [12,30]. Located within the Churchill Province of the Canadian Shield [32], the study region's surface cover primarily consists of glacial and fluvial deposits [31]. Surface waters cover 11–20% of the study region's area, with a quarter of Arctic lakes and ponds characterized as thermokarst landscapes resulting from the loss of ice-rich permafrost [13,15]. Lake Brisson covers an approximate surface area of 45 km²; no depth data is available. Regional data from Kuujjuaq recorded an average total precipitation during sampling years of 490 mm (417–585 mm), of which 42% was as snowfall (166–246 mm) [33].

## 2.2. Field collection

Surface water (n = 63) was collected annually between end-June and mid-September, from 2017 to 2023, and opportunistically in April 2023, amounting to 54 unique sampling stations. Diverse aquatic ecosystems significant to northern regions were sampled: the George River and various tributaries, including the Ford River; the George River estuary; the Koroc River; lakes, including Lake Brisson located adjacent to the Strange Lake REE deposit, and thermokarst lakes (also termed ‘thaw ponds’). A multiparameter meter (*Pro/Pro Plus*, YSI Inc.) was used for in situ surface water quality measurements of temperature, pH, dissolved oxygen (DO), and oxidation-reduction potential (ORP). A clean-hands dirty-hands procedure [38] was considered for surface water collection. From 2 to 4 replicates were acquired for each analysis type at a 30 cm depth with acid-washed (10% hydrochloric acid [HCl]) containers.

Unfiltered samples were taken for total phosphorous (TP), total nitrogen (TN), and total organic carbon (TOC) each year excluding 2018 and 2019. In 2017, filtered samples for dissolved organic carbon (DOC), dissolved inorganic carbon (DIC), and major ions were collected using a 47 mm cellulose acetate filter (0.45 µm pore size) with a filtration tower (Sterifil Aseptic Vacuum Filter System, EMD Millipore) and manual pump. In following years, filtered samples were taken using acid-washed syringes with 0.45µm pore size polyethersulfone (PES) membrane filters (Sarstedt) and included samples for c-DOM analysis in 2019 and 2022. Filtered (F) samples for TMs/REEs were collected using acid-washed syringes (PP/PE) with metal-free 0.45µm pore size PES membrane filters (Whatman, GD/XP). The filters were preconditioned with a few
milliliters of site water before collection of the samples to minimize the
possible retention of metals previously reported [39,40]. We consider
that filtered samples contain truly dissolved species (approx. <0.01µm)
and colloids < 0.45 µm, hereafter collectively referred to as the dis-
solved fraction. Unfiltered (UF) samples for total TMs/REEs were also
collected in 2017, 2022 and 2023. Ultra-trace nitric acid (HNO3; EMD
Millipore) was used as a preservative for all TM/REE samples. Field
blanks were prepared with ultrapure water (Milli-Q; 18.5 ΩMcm) and
handled alongside samples. Containers were double-bagged and stored
until analysis in the fridge (DOC, TOC, TMs/REEs) or freezer (anions,

TP, TN).

## 2.3. Laboratory analysis

Major anions, TP, TN, DIC and DOC samples were sent to Burlington National Lab for Environmental Testing (Environment Canada) in 2017. Major anions (nitrite, nitrate, ammonia, phosphate, bromide, chloride) were quantified by ionic chromatography. As of 2020, TP was analyzed by micro-segmented flow analysis using the colorimetry method 365.3 for phosphorous [41], including persulfate digestion and molybdenum

<figure><img src="image_4.png" /><figcaption>**Fig. 1.** Map of the study area, colored by river basin or sub-watershed, with the stations sampled from 2017 to 2023. A) Map of the permafrost zones of Canada highlighting the study region. B) Zoomed-in view of the Strange Lake complex location. Credits: Esri Canada & Natural Earth Vector 9 [34]; Esri Canada Education and Research (Esri) [35], Natural Resources Canada, & Water Survey of Canada (Environment Canada) [36]. Esri & Garmin; Natural Resources Canada (Government of Canada) [37].</figcaption></figure>
blue complexation (Astoria2 Analyzer, Astoria-Pacific). TN was determined by flow injection analysis (Lachat Quickchem 8500) after persulfate digestion, following colorimetry method 353.2 for nitrogen [42]. TOC and DOC concentrations were quantified using the Aurora 1030 W TOC Analyzer (O.I. Corporation) as of 2018. c-DOM analyses were performed as in Wauthy et al. [15], using a UV-visible Cary 100 spectrophotometer (Agilent Technologies) to measure the absorbance from 250–800 nm. Fluorescent DOM (f-DOM) scans were carried out on a Cary Eclipse fluorescence spectrophotometer (Agilent Technologies) across the excitation waveband of 230–450 nm (5 nm increments) and emission waveband of 240–600 nm (2 nm increments) in order to construct excitation-emission matrices (EEMs). Samples were blanked using ultrapure water (Milli-Q).

Prior to TM/REE analysis, filtered and unfiltered water samples were digested on a hotplate (VWR Scientific) at approximately 100 °C for 2 h [40] with conc. HCl (Omnitrace Ultra, EMD Millipore) and conc. HNO₃ (OmniTrace Ultra in 2017–2019, EMD Millipore; in following years the nitric acid was trace metal grade purified by sub-boiling with Savillex distillation). TMs (²³Na, ²⁴Mg, ²⁷Al, ³⁹K, ⁴⁴Ca, ⁴⁷Ti, ⁵⁵Mn, ⁵⁶Fe, ⁵⁷Fe, ²³²Th, ²³⁸U included herein) and REEs (Y + lanthanides, excluding radioactive Pm) were analyzed by inductively coupled plasma mass spectrometry (ICP-MS/MS, 8900 Triple Quadrupole, Agilent). Certified reference materials SLRS-6 (river water), CALA, and PTC-4 (Table S1–S3), and internal standards ⁴⁵Rh, ⁷⁴Ge, ¹⁸⁵Re, and ¹⁹¹Ir were used for quality control.

## 2.4. Data analysis

Regarding DOM quality, EEMs were corrected in R (Version 3.6.3) for instrument bias, inner filter effect and Raman scattering as described in De Bonville et al. [43], utilizing the R package for PARAFAC correction (Version 0.1) of Labrie et al. [44]. The absorbance at 254 nm was measured, and the slope ratio (S<sub>R</sub>), an index of c-DOM molecular weight, was calculated as in Wauthy et al. [15]. Specific proxies of f-DOM, including the fluorescence index (FI; provides insight into the origin of fulvic acids), the humification index (HIX; indicates degree of humic content), and the freshness (β : α) and biological indices (BIX; both proxies of DOM freshness), were calculated following Hansen et al. [45] and references therein. A PARAFAC model was run on 1190 corrected EEMs and validated by split-half analysis in MATLAB (Version R2013a, MathWorks), following an adapted protocol of Murphy et al. [46]. To identify the main f-DOM components, the excitation and emission spectra were compared to published components available in the OpenFluor database, following Murphy et al. [47]. The PARAFAC model identified four components (Table S4), of which two shared fluorescence characteristics with humic materials from terrestrial or microbial origin (C1 and C2). The third component (C3) displayed a similar source, but was humic- or fulvic-like, while the last component (C4) presented spectra similar to amino acids or proteins, and was attributed to algal production.

For TM/REE analyses, detection limits (DLs) were calculated as three times the standard deviation (SD) of 10 standards (Table S5). Field blank concentrations were subtracted when greater than 15% of sample averages (modified from [48]). The particulate fraction for TMs and REEs was calculated as the difference between unfiltered and filtered concentrations. For principal component analysis (PCA), linear discriminant analysis (LDA) and WHAM modelling only, undetected concentrations were replaced by half their DL. Pearson correlation coefficients demonstrated strong relationships (r = 0.74–1) between concentrations of individual REEs (Table S6). The sum of REEs, or groups of REEs, were therefore presented where relevant. Post-Archean Australian Shale (PAAS) reference values from Pourmand et al. [49] were used for normalization of REEs. Calculations were then performed for cerium anomalies as Ce/Ce* = Ce<sub>N</sub>/(La<sub>N</sub><sup>0.667</sup> • Nd<sub>N</sub><sup>0.333</sup>) [Eq. 1] [50], europium anomalies as Eu/Eu* = Eu<sub>N</sub>/(Sm<sub>N</sub> • Gd<sub>N</sub>)<sup>1/2</sup> [Eq. 2] [51] and gadolinium anomalies as Gd/Gd* = Gd<sub>N</sub>/(0.4Nd<sub>N</sub> + 0.6Dy<sub>N</sub>) [Eq. 3] [52].

Statistical analyses were performed in RStudio (Version 1.3). Variables were log-transformed as needed to improve normality. PCA was performed for complete datasets (n = 48) to select variables that were important contributors of data variance and to minimize overfitting (Table S7). Further refinement of available classed data (n = 54) was done using LDA, which maximized the separation between groups. The selected model underwent k-fold cross-validation with 95% of train and 82% of test predictions accurately classed over 7 iterations (k = 5). Difference of means analyses included Student's t-test and one-way analysis of variance (ANOVA) with Tukey's HSD post-hoc test, using a significance level of 0.05. Linear regression was used for simple and multiple linear models. Analysis of covariance (ANCOVA) and stepwise methods ('step' function) were additionally employed for investigating covariance and for model comparison. Mapping was completed in ArcGIS® Pro (Version 2.4) by ESRI. Speciation modelling was performed on average values with Windermere Humic Aqueous Model (WHAM 7) using 1.0 mg/L DOC as an equivalent for 1.3 mg/L fulvic acid as per Bryan et al. [53], which estimates 50% of DOM is carbon, and 65% of the latter is fulvic acid. DIC concentrations were used as an equivalent for total alkalinity. Default speciation parameters and precipitation of Al (OH)₃ and Fe(OH)₃ were selected. Detailed results were provided in Supplementary Information (Table S8).

# 3. Results and discussion

## 3.1. Water quality of northern aquatic systems

The George and Koroc rivers and their tributaries presented relatively neutral (pH = 6.6–7.0, Table 1) surface waters with average values for DO (9.4–11 mg/L) and ORP (190–230 mV) representative of oxygen-rich freshwater systems. Thermokarst lakes were slightly more acidic, while other lakes were more alkaline, all of which had oxic surface waters. The consistently low nutrient levels across sources (TN = 0.12–0.31 mg/L; TP = 0.0045–0.013 mg/L) were considered characteristic of the region's undeveloped land [54]. Thermokarst lakes and select George River tributaries demonstrated the most organic-rich waters with DOC up to 11.7 mg/L. George River stations extended over 200 km of the river's length, and DOC concentrations remained in the range of 1.8 to 4.6 mg/L. The Koroc River displayed the lowest organic matter content, with both average DOC and TOC < 2 mg/L. These levels were comparable to Arctic freshwater systems, such as the 0.3 to 6.6 mg/L of DOC for Alaskan streams [55]; and large Canadian circumpolar rivers (length > 1000 km), like the Mackenzie River averaging 4.4 ± 4.2 mg/L [56].

Characterization of water sources was performed using dimension reduction techniques. PCA variable contributions (Fig. S1, Table S7) suggested a positive correlation may be present in the study area between the primary metals of interest, the REEs and radionuclides. Also suspected were positive correlations for the REEs with Al and DOC, and to a lesser degree with other factors (i.e., Fe, DO); and a potential negative correlation for REEs with temperature. The selected LDA model (Fig. 2) included 11 variables to explain 89% (LD1:LD2) of variability across groups: the REEs, 4 major cations (Na, Ca, K, Mg), 3 TMs (Al, Fe, U), 1 major anion (Cl⁻), pH and DOC. While an overlap in water quality between George River and its tributaries generated some misclassification among these two groups, this was explained by their connectivity within the GRB. The model consistently classed the Koroc River stations correctly, demonstrating the distinct chemistries of each river basin. The model also correctly classed thermokarst lakes in all cases, highlighting their more isolated nature and the unique chemistry of this landscape feature. However, further permafrost thawing in the subarctic may lead to greater thermokarst lake cover, which could subsequently impact the mobility of metals and nutrients to river systems. Overall, these validated techniques highlighted the range in aquatic ecosystems represented in the study, and the variables critical in driving their geochemical variability.
<table><caption>**Table 1**
Average (SD) values of physicochemical properties, organic matter content and nutrient levels for surface waters in the George and Koroc river basins.</caption><thead><tr><th rowspan="2">Ecosystem</th><th>pH</th><th>DO</th><th>Temp.</th><th>ORP</th><th>DOC</th><th>TOC</th><th>TP</th><th>TN</th></tr><tr><th></th><th>mg/L</th><th>°C</th><th>mV</th><th>mg/L</th><th>mg/L</th><th>µg/L</th><th>mg/L</th></tr></thead><tbody><tr><td>George River</td><td>7.0</td><td>10.6</td><td>12.3</td><td>208</td><td>2.63 (0.72)</td><td>2.27 (0.24)</td><td>4.48 (1.57)</td><td>0.14 (0.05)</td></tr><tr><td>GR-Estuary</td><td>7.6</td><td>11.0</td><td>13.1</td><td>284</td><td>NA</td><td>NA</td><td>4.95 (NA)</td><td>0.12 (NA)</td></tr><tr><td>GR-Tributary</td><td>6.8</td><td>10.6</td><td>11.3</td><td>192</td><td>2.90 (2.33)</td><td>2.12 (0.97)</td><td>4.70 (3.13)</td><td>0.12 (0.03)</td></tr><tr><td>Thermokarst Lakes</td><td>5.9</td><td>10.5</td><td>10.3</td><td>561</td><td>5.72 (3.26)</td><td>5.77 (3.43)</td><td>13.4 (15.1)</td><td>0.31 (0.24)</td></tr><tr><td>Lake Brisson</td><td>7.4</td><td>10.1</td><td>8.5</td><td>201</td><td>1.93 (0.39)</td><td>1.57 (0.02)</td><td>5.12 (0.47)</td><td>0.17 (0.02)</td></tr><tr><td>Unnamed Lake</td><td>7.4</td><td>8.9</td><td>16.1</td><td>266</td><td>2.86 (NA)</td><td>2.82 (NA)</td><td>5.13 (NA)</td><td>0.22 (NA)</td></tr><tr><td>Koroc River</td><td>6.6</td><td>9.4</td><td>15.8</td><td>226</td><td>0.93 (0.55)</td><td>1.61 (1.32)</td><td>6.48 (11.7)</td><td>0.21 (0.20)</td></tr></tbody></table>

<figure><img src="image_5.png" /><figcaption>**Fig. 2.** Biplot of components LD1 and LD2 of the LDA model explaining 89% of variability in surface waters across Nunavik aquatic ecosystems (n = 54) with 11 variables.</figcaption></figure>

## 3.2. REE concentrations in surface waters

Nunavik surface waters had total REE concentrations (Table 2 and S9) within the range of 2.9 to 50.6 nM and dissolved REE concentrations from 2.2 to 46.2 nM (results in µg/L were provided in Table S10). The majority of REEs were in the dissolved fraction (< 0.45 µm) (Table S11), and this trend was even stronger for the HREEs (75–97%) compared to the LREEs (65–93%) by 5% on average (p < 0.05). In general, this finding is consistent with the greater stability of HREEs as dissolved species presented in the literature [57,58]. Further, in consideration of

the known association of REEs to colloids (e.g., [16,17,59,60]), the strong partitioning between fractions suggested a dominance of colloidal REE species (< 0.45 µm) in the study area. In addition, such high recoveries of REEs in the filtered samples indicated that any retention of metals by the filters was likely minimal herein. As this trend was present across all water stations, and there were larger sample sizes (n) for filtered TMs/REEs, discussions primarily focused on the dissolved fraction.

<figure><img src="image_6.png" /><figcaption>**Fig. 3.** Barplot of dissolved REEs, Th and U concentrations (nM) for all sampling years, with stations organized by water source and approximate order within their respective river basins, from upstream (US) to downstream (DS).</figcaption></figure>

<table><caption>**Table 2**
Average (SD) values for concentrations (in nM) of REE groups (total, light, and heavy REEs), U, and Th; and for Ce, Eu, and Gd anomalies in dissolved (<0.45 µm) and total (unfiltered) water fractions. Sample sizes (n) specified for each aquatic ecosystem.</caption><thead><tr><th>Ecosystem<br>DISSOLVED</th><th>n</th><th>REEs<br>nM</th><th>LREEs<br>nM</th><th>HREEs<br>nM</th><th>Ce/Ce*<br>Eq. 1</th><th>Eu/Eu*<br>Eq. 2</th><th>Gd/Gd*<br>Eq. 3</th><th>Th<br>nM</th><th>U<br>nM</th></tr></thead><tbody><tr><td>George River</td><td>21</td><td>4.53 (1.99)</td><td>3.94 (1.81)</td><td>0.59 (0.19)</td><td>0.56 (0.08)</td><td>1.05 (0.14)</td><td>1.03 (0.11)</td><td>0.05 (0.01)</td><td>0.22 (0.08)</td></tr><tr><td>GR-Estuary</td><td>1</td><td>2.18 (NA)</td><td>1.85 (NA)</td><td>0.33 (NA)</td><td>0.46 (NA)</td><td>NA</td><td>NA</td><td>NA</td><td>0.64 (NA)</td></tr><tr><td>GR-Tributary</td><td>17</td><td>10.6 (11.8)</td><td>9.67 (11.1)</td><td>0.90 (0.73)</td><td>0.53 (0.15)</td><td>0.83 (0.21)</td><td>1.07 (0.03)</td><td>0.24 (0.18)</td><td>0.48 (0.43)</td></tr><tr><td>Lake Brisson</td><td>5</td><td>3.13 (1.41)</td><td>2.07 (1.11)</td><td>1.06 (0.62)</td><td>0.68 (0.30)</td><td>0.51 (0.09)</td><td>0.98 (0.03)</td><td>NA</td><td>0.23 (0.12)</td></tr><tr><td>Thermokarst Lakes</td><td>8</td><td>20.1 (15.9)</td><td>17.9 (14.6)</td><td>2.22 (1.50)</td><td>0.89 (0.11)</td><td>0.90 (0.08)</td><td>1.07 (0.02)</td><td>0.45 (0.33)</td><td>0.80 (0.78)</td></tr><tr><td>Unnamed Lake</td><td>1</td><td>0.43 (NA)</td><td>0.36 (NA)</td><td>0.07 (NA)</td><td>0.59 (NA)</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td></tr><tr><td>Koroc River</td><td>10</td><td>2.86 (1.87)</td><td>2.28 (1.50)</td><td>0.58 (0.38)</td><td>0.47 (0.12)</td><td>1.10 (0.21)</td><td>1.12 (0.10)</td><td>NA</td><td>0.08 (0.03)</td></tr><tr><td><strong>TOTAL</strong></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>George River</td><td>5</td><td>6.70 (2.06)</td><td>5.96 (1.96)</td><td>0.74 (0.11)</td><td>0.61 (0.05)</td><td>NA</td><td>1.19 (0.02)</td><td>NA</td><td>0.32 (0.11)</td></tr><tr><td>GR-Estuary</td><td>1</td><td>4.46 (NA)</td><td>3.89 (NA)</td><td>0.57 (NA)</td><td>0.61 (NA)</td><td>NA</td><td>1.19 (NA)</td><td>NA</td><td>0.83 (NA)</td></tr><tr><td>GR-Tributary</td><td>3</td><td>18.0 (13.7)</td><td>16.8 (13.0)</td><td>1.24 (0.74)</td><td>0.66 (0.05)</td><td>NA</td><td>1.14 (0.06)</td><td>NA</td><td>0.40 (0.26)</td></tr><tr><td>Lake Brisson</td><td>2</td><td>3.20 (0.39)</td><td>1.76 (0.06)</td><td>1.44 (0.45)</td><td>0.56 (0.05)</td><td>0.50 (0.04)</td><td>0.97 (0.00)</td><td>NA</td><td>0.36 (0.09)</td></tr><tr><td>Thermokarst Lakes</td><td>8</td><td>21.7 (16.6)</td><td>19.3 (15.2)</td><td>2.47 (1.58)</td><td>0.89 (0.11)</td><td>0.91 (0.05)</td><td>1.05 (0.02)</td><td>0.43 (0.31)</td><td>0.91 (0.87)</td></tr><tr><td>Unnamed Lake</td><td>1</td><td>0.75 (NA)</td><td>0.64 (NA)</td><td>0.11 (NA)</td><td>0.66 (NA)</td><td>NA</td><td>NA</td><td>NA</td><td>NA</td></tr><tr><td>Koroc River</td><td>4</td><td>6.11 (1.36)</td><td>4.90 (1.09)</td><td>1.21 (0.28)</td><td>0.45 (0.10)</td><td>0.89 (0.02)</td><td>1.16 (0.04)</td><td>NA</td><td>0.13 (0.03)</td></tr></tbody></table>
The levels of REEs in the river mainstems were relatively stable compared to the variability seen between tributaries (Fig. 3). For example, the Ford River tributary (Table S10) had average (± SD) dissolved REE concentrations of 3.2± 1.3 nM, whereas two unnamed tributaries consistently displayed REE levels an order of magnitude greater. Similar levels were reported for undisturbed environments where REEs ranged from 0.09 to 20.8 nM in Eastern Canadian lakes and 0.6 to 24.4 nM (excluding Y, Tb) in various Arctic Siberian rivers [48, 61]. Case studies for industrially impacted waters have reported REE concentrations orders of magnitude greater than those presented herein for Nunavik ecosystems (e.g., [62]). Overall, due to the relatively low concentrations of REEs, in addition to the lack of anthropogenic influence in the study region, results were interpreted as background concentrations representative of undisturbed surface waters. Environmental pressures, stemming from climate change and mining exploration, will increasingly and collectively impact these fragile northern ecosystems. We therefore provide detailed baselines values for individual REEs (Table S10) as a reference for future studies investigating potential changes in the region.

A LREE enrichment was seen in Nunavik waters, with higher dissolved concentrations of LREEs as compared to HREEs by up to 20-times (LREEs/HREEs = 1–20). In addition to the greater abundance of LREEs in the environment, the rivers’ slightly acidic to circumneutral waters may help stabilize the LREEs in the water column, as they are more strongly removed by colloids at alkaline conditions [63,64]. In fact, this trend corresponds with a low to absent enrichment of LREEs in the more alkaline Lake Brisson waters. A previous study in the GRB demonstrated LREE enrichments (LREEs/HREEs = 2–70) for sediments and biota (various species) of different ecosystems [65]. Together, this supports that a natural LREE enrichment is prevalent across the study region.

The aquatic speciation of REEs is also an important driver of their distribution and must be considered when discussing the effects of REEs on aquatic organisms. The REEs can bioaccumulate in plants and animals, though no biomagnification along the trophic chain is suspected [65]. The REE species have demonstrated differences in bioavailability, with the free metal species (Ln³⁺) considered a major bioavailable form [4100] and good predictor of bioaccumulation in freshwater zooplankton [48]. REE exposure studies have demonstrated toxic effects on aquatic species, for example, with LC₅₀ values on the order of 10² nM for freshwater crustaceans (*D. pulex*, *H. azteca*) [66]. Additionally, some evidence of higher toxicity for HREE species has been reported from lab studies [67]. Due to the very low solubility of the REEs, they typically only occur to a small degree as the free ion species (REE³⁺) in freshwater waters [18]. Speciation modelling (WHAM) of Nunavik waters in the current study (Table S8) predicted that free ions accounted for between 0.73% of La in thermokarst lakes, to < 10⁻⁴% of Yb in Lake Brisson. While these represent small concentrations (up to 10⁻² nM), chronic low exposure may lead to toxic effects, and the bioavailable fraction could be underpredicted if other species are also accumulated by organisms as REE-complexes [68]. Anthropogenic activities reportedly led to a three-fold increase in REE solubility compared to non-mining regions, signifying important implications for their bioavailability [69]. As speciation may be sensitive to even minor changes in water quality, there is a risk to ecosystem health with ongoing environmental pressures.

### 3.3. Influence of water temperature on REEs

A main environmental influence on dissolved REE concentrations was demonstrated by a negative log-linear relationship to surface water temperature (R² = 0.23, p < 0.001) (Fig. 4). The model (slope = -0.045) represented a decrease in REE concentrations on the order of 10¹ with increasing temperatures over the measured range of 2.4–19.5 °C. This approximated to a predicted change of 10% in REE concentration per degree of change in temperature, according to the model equation. Further, the regressions were similar between REE groups (slopes =

<figure><img src="image_7.png" /><figcaption>Fig. 4. Linear regressions (p < 0.01) of dissolved REEs with in situ surface water temperatures (°C) in Nunavik rivers, streams, alkaline lakes, and thermokarst lakes. The LREEs and HREEs are represented by La and Yb, respectively.</figcaption></figure>

-0.04), though a stronger correlation was presented for Yb (R² = 0.34, p < 0.001) than for La (R² = 0.18, p < 0.01), representing the HREEs and LREEs respectively. In addition, the correlation strengths (R²) were improved to 0.41 (Yb) and 0.40 (La) when the regression was modelled with river data only.

Exploration of the temperature relationship in natural surface waters was not found in the literature. However, the current study covered a reasonably large range of surface water temperatures that occurs in the northern environment, offering unique insight into REE trends. Water temperature may covary with other environmental factors, such as water pH, salinity, and DO content. However, the presented relationship for REEs was not simply explained by an indirect influence on temperature by pH or DO, which had negligible contributions to model fitness (ANCOVA) and were excluded by stepwise regression (Table S12). Chloride (Cl⁻) improved the model as an additional independent variable, noting that the covariance term was insignificant. These findings were consistent with the weak correlations (Fig. S2) for dissolved REEs with pH (R² = 0.08) or DO (R² = 0.14), and the more pronounced log-log relationship with Cl⁻ concentrations (R² = 0.26). Further, the expected linear correlation of temperature with DO was strong (R² = 0.67, p < 0.001; Fig. S3), but the temperature was otherwise unrelated to tested environmental factors (pH, DOC, Cl⁻; Fig. S3). Speciation simulations (WHAM) supported this relationship to temperature (Fig. S4), which may be related to decreasing pCO₂ in warming waters resulting in solubility changes [70-72], such as with lower REE-carbonates in solution (Fig. S5) [58]. Though no covariance was identified with pH herein, the impact of changing pCO₂ on the carbonate equilibrium and consequently, the alkalinity [70], further demonstrates the complex interplay of multiple environmental factors. According to interaction modelling of Millero et al. [73], the REEs are among the metals expected to be the most significantly impacted by climate change in seawaters due to the important formation of REE-carbonate complexes. Future investigations should aim at better understanding this phenomenon and refining predictive models, as field data suggests that warming in the North may influence cycling of REEs in climate-impacted systems, such as by promoting a precipitation of REEs out of dissolution and further enhancing LREE enrichments in surface waters.

While pH did not serve as an important factor driving REE concentrations (Fig. S2), the relatively negative trend is consistent with previous studies in river systems, in which lower aquatic REE concentrations were present at higher pH values [64,74]. In the current study, the highest REE concentrations were found in thermokarst lakes with more acidic waters (5.5–6.5). The effect of pH on REE distribution has elsewhere been attributed to changes in particle surface reactions and the solubility constants of dissolved complexes [75,76]. The pH of a system may also contribute through an indirect influence on ligand availability, such as the presence of organic matter [17]. However, the
pH was not significantly related to the DOC concentration in the present study ($p > 0.05$). Further, the increase in alkalinity and salinity as the river mixes in with the Ungava Bay waters was expected to contribute to a decrease in REE solubility [75]. Though caution is necessary when making interpretations of the estuarian results due to limited sampling ($n = 1$), data suggested a loss of REEs in the estuary by 56% and 35% for the dissolved and total fractions, respectively. The non-conservative mixing of REEs is likely attributed to the strong coagulation of colloids in the mixing zone, which along with the adsorption of particle-reactive species, causes precipitation of metals out of the water column [22,77].

### 3.4. Radionuclides (U, Th) and their Co-occurrence with REEs

The measured radionuclides of thorium ($^{232}$Th) and uranium ($^{238}$U) represent their most abundant isotopes ($\ge 99\%$), which have long half-lives of 14 By and 4.47 By, respectively, and produce the less stable nuclides $^{228}$Ra and $^{234}$Th through alpha decay [78]. Half of the water stations in the study region had concentrations of Th (Fig. 3) below the DL (average DL: 0.002 nM; Table S5), which may be related to the very poor solubility of free Th in circumneutral waters and its low mobility from refractory minerals [79]. Uranium concentrations (Fig. 3) were frequently detected and remained low ($< 3$ nM), consistent with the reported values of $< 16.8$ nM for rivers without localized enrichments [80]. In oxic waters, uranium is likely present in the U(VI) form as soluble and mobile uranyl ions ($UO_2^{2+}$), which have the potential to complex with various anions (e.g., carbonates, hydroxides), inorganic ligands (e.g., metal oxides, clays) and the humic substances of DOM [81, 8, 80]. Higher concentrations of U have been reported in surface and groundwaters near uranium deposits and in contaminated sites such as those in proximity to mining or industrial processes [80]. The Canadian guideline for total uranium in drinking waters of 20 $\mu$g/L (approximately 80 nM) [82] further highlights the naturally low concentrations of U reported herein.

In the present study, radionuclides were significantly correlated ($p < 0.001$) to concentrations of REEs in filtered surface waters (Fig. 5), with positive log-log relationships for Th ($R^2 = 0.79$) and U ($R^2 = 0.27$).

While both river basins are within the Churchill geological province of the Canadian Shield [32], the GRB presented elevated radionuclide ($p < 0.01$) and REE concentrations ($p < 0.05$) relative to the KRB. This trend may be related to differences in their local geological features, such as the presence of REEs-Th-U co-mineralization in the GRB [6,83]. Additionally, while both Th and U are in the actinide series, differences in their reactivity during water-rock interactions may contribute to the decoupling of their distribution in dissolved surface waters (lm: $p > 0.05$). In particular, uranium has a greater potential for leaching and mobilization during weathering from parent rocks to soils, rivers and lakes [81,84,79]. The differences in geochemical behavior between radionuclides likely contributed to the stronger correlation of Th with the similarly insoluble and chemically resistant REEs, as compared to U, which was not suggestive of a co-transport with REEs in surface waters.

### 3.5. Drivers of REE speciation and transport

#### 3.5.1. Trace metal complexation

Significant, positive log-log relationships (Fig. 5) with Al ($R^2 = 0.57$) and Fe ($R^2 = 0.52$) were reported for REEs in filtered surface waters within the study region. This relationship was not significant for the particulate fraction. Due to their multiple redox states (e.g., Fe[0], Fe[II], Fe[III]), these elements are sensitive to the surrounding redox conditions, which influences their solubilities and affinities for ligands [85]. Given the high DO content of the surface waters (Table 1), elements are presumably in their oxidized form, such as the more insoluble $Fe^{3+}$ oxide and hydroxide species. The presented relationships may indicate the adsorption of REEs to inorganic (oxyhydr)oxide colloids of Fe and Al, which have previously been reported as important drivers of REEs in river waters [86-88]. With such high correlations to Al, a presence of REE aluminosilicate colloids may also contribute to the transport of REEs, such as fine clay particles [25]. To note, competition between the REEs and TMs ($Al^{3+}$, $Fe^{3+}$), such as for functional sites on DOM, have been discussed in the literature [89]. However, this is not considered a major driver within the present study as the trends displayed positive relationships without a plateau at higher concentrations, likely due to the near-neutral pH levels favoring the dissolution of TM

<figure><img src="image_8.png" /><figcaption>**Fig. 5.** Relationships in dissolved (D) surface waters for REEs (nM) with DOC (mg/L), highly correlated trace metals (Al, Fe; nM), and radionuclides (U, Th; nM). Relationships in the particulate fraction (P) were also presented for REEs with Al and Fe. Linear regressions displayed for significant models (dissolved fractions only).</figcaption></figure>
competitors [24]. Instead, the correlations presented herein may also be consistent with the potential co-transport of REEs, Fe, and Al by organic colloids, and may signify organo-Al as a main carrier [61].

### 3.5.2. Organic matter complexation

Significant positive relationships were presented for REEs with DOC concentrations (R² = 0.40, *p* < 0.001) in the dissolved pool. A similar relationship for REEs with TOC (R² = 0.40, *p* < 0.001) was simply attributed to the strong DOC control on organic matter content, as DOC represented ≥95% of the TOC in summer months, where both were analyzed. The importance of organic REE species has been presented for natural waters [90,91] and is further supported by speciation modelling (WHAM), which predicted that over 99% of dissolved REEs in each ecosystem were associated with organic matter, as fulvic acids (Table S8).

In addition to DOM content, its quality may also influence metal speciation in surface waters [92], explored through PARAFAC modelling of c-DOM [93]. Log-log relationships between REEs and the four carbon components (C1–C4) were made for filtered waters (Fig. 6 and S6). The humic- and fulvic-like carbon components were the strongest explanatory variables for REE distribution with C2 (R² = 0.85); and to a lesser extent both C1 and C3 (R² = 0.31). The model for C4 was not significant (*p* > 0.05), suggesting that autochthonous carbon production within the water column (e.g., in situ microbial activity) was not an important contributor to the fate of REEs in these subarctic waters. These relationships were supported by trends (Fig. S7) with the consistently low BIX values (0.45–0.58; Table S13) and freshness indices (β : *α* = 0.43–0.56), and intermediate to high HIX values (3–14) across GRB waters [45,94]. In further support, dissolved REEs within the St-Lawrence River basin (QC, Canada) were also explained by strong, positive linear models with humic-like carbon components (R² = 0.52–0.87), and only weak, negative linear models were present for the tyrosine-like component [40]. Altogether, the model predicted that the organic mobilization of REEs in the GRB was primarily driven by humic-rich, less fresh carbon (not microbial-derived) with a terrestrial-dominated source. As increased mobilization of organic matter from terrestrial environments reportedly occurs with climate change [56,95,96], the relationships suggest greater loads of continental REEs may be delivered to aquatic ecosystems with further surface warming.

The quality of organic matter in natural waters is an important factor controlling the toxicity of metals to organisms, linked to the strong affinity of humic substances for metals (e.g., Cd, Cu, Ni, Zn) [97] and REEs

[40]. Multiple studies on aquatic systems have demonstrated that DOM influences metal bioavailability and bioaccumulation [98,48,68]. MacMillan et al. [48] conducted field sampling of Canadian lakes that demonstrated the REE:DOC ratio was an important predictor in the concentration of REEs in freshwater zooplankton, where DOC seemingly lowered the degree of individuals’ REE bioaccumulation. The forecasted transformations of organic matter loads and composition along the permafrost gradient in response to climate change in the subarctic environment [99,13,15] may therefore alter REE speciation, consequently impacting the bioavailability of REEs to aquatic organisms with uncertain ecotoxicological implications.

### 3.6. Anomalies of Ce, Eu and Gd

The river waters had significant negative Ce anomalies (Ce/Ce* = 0.18–0.79, Table 2 and S9), and less pronounced Ce anomalies were present in thermokarst lakes. In surface environments, cerium is oxidized to the more insoluble and reactive Ce⁴⁺ that tends to naturally precipitate with colloids or particulates in the water column [26]. The Ce anomalies presented for Nunavik river waters were therefore considered representative of the naturally oxidizing environments (ORP > 0 mV, i.e., redox potential) [100]. The relationships between Ce anomalies and concentrations of dissolved TMs were studied in Nunavik surface waters, with positive models (Fig. S8) presented for Fe (R² = 0.52, *p* < 0.001), Al (R² = 0.39, *p* < 0.001) and a novel relationship to Ti (R² = 0.28, *p* < 0.01). The correlation of these variables may be linked to the occurrence of redox reactions between Ce and multi-valent TM colloids [101,102]. This process is termed oxidative scavenging and typically associates the oxidation and adsorption of Ce with Fe or Mn-(oxyhydr)oxides [22,27]. No tested models for REEs distribution (e. g., Tables S12 and S14) or Ce anomaly formation were improved by considering Mn concentrations of Nunavik surface waters. Further, as Ti and Al may enter surface waters through the transport of weathered soils or bedrock [103], the link may reflect an allochthonous source for these colloidal oxides. Additionally, the diminution in magnitude of the Ce anomaly with increasing TM concentrations may be an indirect representation of organic matter influence, which has a reported effect on Ce fractionation [16,102]. Further investigation revealed significant positive relationships for Ce/Ce* with DOC concentrations (Fig. S8) and PARAFAC parameters (Fig. S9), particularly a positive correlation with HIX (R² = 0.83, *p* < 0.001). The inorganic colloids may therefore be of greater importance in low-DOC systems, like the Koroc River, that preferentially take up Ce, producing more negative Ce anomalies; and of

<figure><img src="image_9.png" /><figcaption>**Fig. 6.** c-DOM components (PARAFAC) as explanatory factors for dissolved REE concentrations in GRB surface waters. Regression line and statistics displayed for significant models (solid: both years; dashed: 2019 only).</figcaption></figure>
lesser control in DOC-rich waters, like thermokarst lakes, which have an affinity for all REEs. The presence of a variety of reactive functional groups (e.g., phenols) within humic substances allows for their partici- pation in redox reactions with TMs [104]. We propose a redox control on REE fractionation that may arise from Ce interaction with Fe/Al-(ox- yhydr)oxides and organic colloids, with relative influences that may rely on factors such as ligand content and water chemistry.

Surface waters of the river systems presented near-unit mean Eu anomalies (Eu/Eu* = 0.83–1.10, Table 2). Europium is typically in the trivalent state in oxidizing surface environments; while its reduction to Eu²⁺ is favored under anoxic and/or high-temperature conditions [105]. Europium anomalies can be induced by the weathering of terrains enriched in minerals with Eu substitutions (e.g., feldspars), formed during magmatic or hydrothermal processes [106,63,107]. The negative Eu anomalies presented at Lake Brisson (Eu/Eu* = 0.51 ± 0.09) may be reflective of the local geology of the Strange Lake complex, which hosts Eu-depleted granites and pegmatites [6]. A slightly stronger Eu frac- tionation was reported for lakes and rivers from Kuujjuarapik (Nunavik) with average anomaly values of 1.9 [108]. However, Eu/Eu* values presented herein are within the range typically presented as background values for oxygen-rich waters (e.g., [107,109]). Additionally, the stud- ied water bodies displayed weak to absent Gd/Gd* anomalies, with mean values of 0.98–1.12 across ecosystems (Table 2). The Gd anomaly has elsewhere been used as an indicator of human activity within a watershed, with large positive anomalies (≥ 10) suggesting nearby REE sources such as hospital effluents or city wastewaters (e.g., due to Gd in contrast agents for MRIs) [40]. The absence of significant positive Gd anomalies (e.g., above 1.8) [23] supports that no important human ac- tivities were detectable in the region, which was expected as the study area is primarily undeveloped, with a population of 951 for the com- munity of Kangiqsualujjuaq [110].

### 3.7. Implications for northern regions

Worldwide, REEs have been studied to a greater extent in contami- nated regions than undisturbed locations, with limited studies that focus on northern freshwaters and an underrepresentation of Canadian eco- systems in REE literature [111]. The provided baseline data supports future research regarding environmental changes in subarctic regions, serving as a reference for studies investigating climate change effects and impacts of anthropogenic activities. In addition, the multi-annual field sampling contributed to the empowerment of an Inuit commu- nity facing potential mining exploitation on its territory, which is an important source of traditional food for community members.

### Environmental implications

Rare earth elements (REEs) and co-occurring radionuclides present environmental risks to aquatic systems and were investigated in sub- arctic surface waters (rivers, lakes) before mining exploitation. REEs and Ce fractionation were driven by Al, Fe, Ti, U, Th and DOC, suggesting a redox control and potential co-transport. Novel correlations with water temperature and mobilization by terrestrial colloids highlighted changes to REE geochemistry and consequentially, bioavailability, with continued climate warming.

### Declaration of funding

The project was funded through a Master research scholarship of the Fonds de recherche Nature et technologies Québec (FRQNT); the Northern Scientific Training Program, Polar Knowledge Canada; and the Program for Collaborative Partnership Projects, Groupe de recherche interuniversitaire en limnologie (GRIL) [H.M.]. Additional support was provided by the Canada Research Chairs Program [M.A.]. The research was also funded through the Northern Scientific Training Program; the Alexander Graham Bell Canada Scholarship (NSERC); and the W.

Garfield Weston Award for Northern Research [G.M.]. Further support was received through: ArcticNet, a Network of Centres of Excellence of Canada [Project # 21 (2019–2023)]; the Crown-Indigenous Relations and Northern Affairs Canada (CIRNAC) [Grant # 1920-HQ-000150]; Arctic Inspiration Prize; and Polar Knowledge Canada (PKA-1718-0012; PKA-2223-0057).

### 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.

### Data Availability

Data will be made available on request.

### Acknowledgements

Authors would like to thank M. Chrifi-Alaoui and D. Bélanger for their laboratory analyses and continued support. Thank you to the members of the community of Kangiqsualujjuaq for their collaboration, through which the study originated. Additionally, we acknowledge that the activities of the Imalirijiit Science and Culture camps facilitated field sampling at select locations. Thank you to Nunavik Parks and the Kativik Regional Government for their contributions to field sampling. We would also like to thank T. Groleau and M. St-Aubin for their assistance in the field.

## Appendix A. Supporting information

Supplementary data associated with this article can be found in the online version at doi:10.1016/j.jhazmat.2024.134418.

### References

[1] Gunn, G., 2014. Critical Metals Handbook. American Geophysical Union, Wiley, (http://ebookcentral.proquest.com/lib/umontreal-ebooks/detail.action?doc ID=1594545).

[2] Wang, J.L., Guo, M.Y., Liu, M.M., Wei, X.Q., 2020. Long-term outlook for global rare earth production. Resour Policy 65. https://doi.org/10.1016/j. resourpol.2019.101569.

[3] Arienzo, M., Ferrara, L., Trifuoggi, M., Toscanesi, M., 2022. Advances in the fate of rare earth elements, REE, in transitional environments: coasts and estuaries. Water 14 (3). https://doi.org/10.3390/w14030401.

[4] Jowitt, S.M., Werner, T.T., Weng, Z.H., Mudd, G.M., 2018. Recycling of the rare earth elements. Curr Opin Green Sustain Chem 13, 1–7. https://doi.org/10.1016/j.cogsc.2018.02.008.

[5] Batley, G.E., Campbell, P.G.C., 2022. Metal contaminants of emerging concern in aquatic systems. Environ Chem 19 (1), 23–40. https://doi.org/10.1071/ EN22030.

[6] Kerr, A., & Rafuse, H. (2012). *Rare-Earth Element (REE) Geochemistry of the Strange Lake Deposits: Implications for Resource Estimation and Metallogenic Models*. Geological Survey, Newfoundland and Labrador Department of Natural Resources Retrieved from https://www.gov.nl.ca/iet/files/mines-geoscience- publications-currentresearch-2012-kerr-rafuse-2012.pdf.

[7] Ma, Y.H., Wang, J.K., Peng, C., Ding, Y.Y., He, X., Zhang, P., Li, N., Lan, T., Wang, D.Q., Zhang, Z.H., Sun, F.H., Liao, H.Q., Zhang, Z.Y., 2016. Toxicity of cerium and thorium on *Daphnia magna*. Ecotoxicology Environ Saf 134, 226–232. https://doi.org/10.1016/j.ecoenv.2016.09.006.

[8] Markich, S.J., 2002. Uranium speciation and bioavailability in aquatic systems: an overview. Sci World J 2, 707–729. https://doi.org/10.1100/tsw.2002.130.

[9] Yin, X.B., Martineau, C., Demers, I., Basiliko, N., Fenton, N.J., 2021. The potential environmental risks associated with the development of rare earth element production in Canada. Environ Rev 29 (3), 354–377. https://doi.org/10.1139/er- 2020-0115.

[10] Rantanen, M., Karpechko, A.Y., Lipponen, A., Nordling, K., Hyvärinen, O., Ruosteenoja, K., Vihma, T., Laaksonen, A., 2022. The Arctic has warmed nearly four times faster than the globe since 1979. Commun Earth Environ 3 (1). https:// doi.org/10.1038/s43247-022-00498-3.

[11] Allard, M., Lemay, M., Barrette, C., L'Hérault, E., Sarrazin, D., Bell, T., Doré, G., 2012. Permafrost and climate change in Nunavik and Nunatsiavut: Importance for municipal and transportation infrastructures. In: Allard, M., Lemay, M. (Eds.), Nunavik and Nunatsiavut: From Science to Policy - An Integrated Regional Impact Study (IRIS) of Climate Change and Modernization. ArcticNet Inc, p. 303.
[12] Bayle, A., Roy, A., Dedieu, J.P., Boudreau, S., Choler, P., Lévesque, E., 2022. Two distinct waves of greening in northeastern Canada: summer warming does not tell the whole story. *Environ Res Lett* 17 (6). https://doi.org/10.1088/1748-9326/ac74d6.

[13] Vonk, J.E., Tank, S.E., Bowden, W.B., Laurion, I., Vincent, W.F., Alekseychik, P., Amyot, M., Billet, M.F., Canario, J., Cory, R.M., Deshpande, B.N., Helbig, M., Jammet, M., Karlsson, J., Larouche, J., MacMillan, G., Rautio, P., Anthony, K.M. W., Wickland, K.P., 2015. Reviews and syntheses: Effects of permafrost thaw on Arctic aquatic ecosystems. *Biogeosciences* 12 (23), 7129–7167. https://doi.org/10.5194/gb-12-7129-2015.

[14] Pourret, O., Davranche, M., Gruau, G., Dia, A., 2007. Rare earth elements complexation with humic acid. Chem Geol 243 (1-2), 128–141. https://doi.org/10.1016/j.chemgeo.2007.05.018.

[15] Wauthy, M., Rautio, M., Christoffersen, K.S., Forsström, L., Laurion, I., Mariash, H.L., Peura, S., Vincent, W.F., 2018. Increasing dominance of terrigenous organic matter in circumpolar freshwater s due to permafrost thaw. *Limnol Oceanogr Lett* 3 (3), 186–198. https://doi.org/10.1002/lo12.10063.

[16] Dia, A., Gruau, G., Olivie-Lauquet, G., Riou, C., Molenat, J., Curmi, P., 2000. The distribution of rare earth elements in groundwaters: Assessing the role of source-rock composition, redox changes and colloidal particles. *Geochim Et Cosmochim Acta* 64 (24), 4131–4151. https://doi.org/10.1016/S0016-7037(00)00494-4.

[17] Ingri, J., Widerlund, A., Land, M., Gustafsson, O., Andersson, M., Ohlander, B., 2000. Temporal variations in the fractionation of the rare earth elements in a boreal river; the role of colloidal particles. Chem Geol 166 (1-2), 23–45. https://doi.org/10.1016/S0009-2541(99)00178-3.

[18] Sneller, F.E.C., Kalf, D.F., Weltje, L., & Van Wezel, A.P. (2000). Maximum Permissible Concentrations and Negligible Concentrations for Rare Earth Elements (REEs) (RIVM report 601501 011). National Institute of Public Health and the Environment.

[19] Davranche, M., Gruau, G., Dia, A., Marsac, R., Pedrot, M., Pourret, O., 2015. Biogeochemical factors affecting rare earth element distribution in shallow wetland groundwater. *Aquat Geochim 21* (2-4), 197–215. https://doi.org/10.1007/s10498-014-9247-6.

[20] Garcia, M.G., Lecomte, K.L., Pasquini, A.I., Formica, S.M., Depetris, P.J., 2007. Sources of dissolved REE in mountainous streams draining granitic rocks, Sierras Pampeanas (Cordoba, Argentina). *Geochim Et Cosmochim Acta* 71 (22), 5355–5368. https://doi.org/10.1016/j.gca.2007.09.017.

[21] Censi, P., Sposito, F., Inguaggiato, C., Zuddas, P., Inguaggiato, S., Venturi, M., 2018. Zr, Hf and REE distribution in river water under different ionic strength conditions. *Sci Total Environ* 645, 837–853. https://doi.org/10.1016/j.scitotenv.2018.07.081.

[22] Pourret, O., Tuduri, J., 2017. Continental shelves as potential resource of rare earth elements. Nat: Sci Rep 7, 5857. https://doi.org/10.1038/s41598-017-06380-z.

[23] Louis, P., Messaoudene, A., Jrad, H., Abdoul-Hamid, B.A., Vignati, D.A.L., Pons, M.N., 2020. Understanding Rare Earth Elements concentrations, anomalies and fluxes at the river basin scale: The Moselle River (France) as a case study. *Sci Total Environ* 742, https://doi.org/10.1016/j.scitotenv.2020.140619.

[24] Marsac, R., Catrouillet, C., Davranche, M., Bouhnik-Le Coz, M., Briant, N., Janot, N., Otero-Fariña, A., Groenenberg, J.E., Pédrot, M., Dia, A., 2021. Modeling rare earth elements binding to humic acids with model VII. Chem Geol 567. https://doi.org/10.1016/j.chemgeo.2021.120099.

[25] Steinmann, M., Stille, P., 2008. Controls on transport and fractionation of the rare earth elements in stream water of a mixed basaltic-granitic catchment basin (Massif Central, France). Chem Geol 254 (1-2), 1–18. https://doi.org/10.1016/j.chemgeo.2008.04.004.

[26] Tostevin, R. (2021). *Cerium Anomalies and Paleoredox*. Cambridge University Press. https://doi.org/10.1017/9781108847223.

[27] Bau, M., 1999. Scavenging of dissolved yttrium and rare earths by precipitating iron oxyhydroxide: Experimental evidence for Ce oxidation, Y-Ho fractionation, and lanthanide tetrad effect. *Geochim Et Cosmochim Acta* 63 (1), 67–77. https://doi.org/10.1016/S0016-7037(99)00014-9.

[28] Tornvat. (2023). Tornvat Metals. Retrieved 2023-07-07 from https://tornvatmetals.com.

[29] Gérin-Lajoie, J., Herrmann, T.M., MacMillan, G.A., Hébert-Houle, É., Monfette, M.G., Rowell, J.A., Anaviapik Soucie, T., Snowball, H., Townley, E., Lévesque, E., Amyot, M., Franssen, J., Dedieu, J.-P., 2018. IMALIRIJIT: a community-based environmental monitoring program in the George River watershed, Nunavik, Canada. *Écience* 25 (4), 381–399. https://doi.org/10.1080/11956860.2018.1498226.

[30] Kativik Regional Government. (2005). *Kuururjuaq Park Project* (Monts-Tornvat-et-Rivière Koroc). *Status Report*. Kuujjuaq, Québec: Kativik Regional Government (KRG).

[31] Sicaud, E., Fortier, D., Dedieu, J.P., Franssen, J., 2024. Pairing remote sensing and clustering in landscape hydrology for large-scale changes identification: an application to the subarctic watershed of the George River (Nunavik, Canada). *Hydrol Earth Syst Sci* 28, 65–86. https://doi.org/10.5194/hess-28-65-2024.

[32] Ministère des Ressources naturelles et des Forêts. (2023). *Online products and services - Mines, Interactive Map* [Interactive map]. Ministère des Ressources naturelles et des Forêts (MRNF). (https://sigeom.mines.gouv.qc.ca/signet/classes/11108_afchCartelntr).

[33] weatherstats.ca based on Environment and Climate Change Canada data. (n.d.). *Past weather and climate - Historical Data Canada Weather Stats* (weatherstats.ca). (https://weather.gc.ca).

[34] Esri Canada; Natural Earth Vector (2013). Provinces and Territories of Canada. [Layer] https://www.arcgis.com/home/item.html?id=d3fe65386df4e63b02d6e23bb98a1ee.

[35] Esri Canada Education and Research, Esri; Natural Resources Canada; Water Survey of Canada, Environment Canada. (2015). Watersheds in Canada. [Layer] https://hub.arcgis.com/maps/12b6e33d5a754c92b97ae5d0fed9640a/about.

[36] Esri, Garmin. (2011). World Water Bodies. [Layer] https://www.arcgis.com/home/item.html?id=e750071279bf450cbd510454a80f2e63.

[37] Natural Resources Canada, Government of Canada. (2016). Permafrost Atlas of Canada, 5th Edition. [Shapefile] Atlas of Canada, 5th Edition. https://open.canada.ca/data/en/dataset/d1e2048b-cff-5852-aaa5-b861bd55c367.

[38] St. Louis, V.L., Rudd, J.W.M., Kelly, C.A., Beaty, K.G., Bloom, N.S., Flett, R.J., 1994. Importance of Wetlands as Sources of Methyl Mercury to Boreal Forest Ecosystems. Can J Fish Aquat Sci 51 (5), 1065–1076. https://doi.org/10.1139/f94-106.

[39] Jreije, I., Hadioui, M., Wilkinson, K.J., 2022. Sample preparation for the analysis of nanoparticles in natural waters by single particle ICP-MS. Talanta 238. https://doi.org/10.1016/j.talanta.2021.123060.

[40] Lafrenière, M.C., Lapierre, J.F., Ponton, D.E., Guillemette, F., Amyot, M., 2023. Rare earth elements (REEs) behavior in a large river across a geological and anthropogenic gradient. *Geochim Et Cosmochim Acta* 353, 129–141. https://doi.org/10.1016/j.gca.2023.05.019.

[41] U.S. Environmental Protection Agency (EPA), National Exposure Research Laboratory (NERL), (1978). EPA 365.3: Phosphorous, All Forms (Colorimetric, Ascorbic Acid, Two Reagent). (https://www.nemi.gov/methods/method_summaryy/5255/).

[42] U.S. Environmental Protection Agency (EPA), National Exposure Research Laboratory (NERL), (1992). EPA 353.2: Nitrogen, Nitrate-Nitrite (Colorimetric, Automated, Cadmium Reduction). (https://www.nemi.gov/methods/met hod_summary/4702/).

[43] De Bonville, J., Amyot, M., del Giorgio, P., Tremblay, A., Bilodeau, F., Ponton, D. E., Lapierre, J.F., 2020. Mobilization and transformation of mercury across a dammed boreal river are linked to carbon processing and hydrology. Water Resour Res 56 (10). https://doi.org/10.1029/2020WR027951.

[44] Labrie, R., Fortin St-Gelais, N., & Bélanger, S. (2017). *paRafac.correction: an R Package to correct EEMs and derive CDOM and FDOM indices*. In (Version 0.1) https://github.com/RichardLaBrie/paRafac.correction/.

[45] Hansen, A.M., Kraus, T.E.C., Pellerin, B.A., Fleck, J.A., Downing, B.D., Bergamaschi, B.A., 2016. Optical properties of dissolved organic matter (DOM): Effects of biological and photolytic degradation. *Limnol Oceanogr* 61 (3), 1015–1032. https://doi.org/10.1002/lno.10270.

[46] Murphy, K.R., Stedmon, C.A., Graeber, D., Bro, R., 2013. Fluorescence spectroscopy and multi-way techniques. *PARAFAC*. Anal Methods 5 (23), 6557–6566. https://doi.org/10.1039/c3ay41160e.

[47] Murphy, K.R., Stedmon, C.A., Wenig, P., Bro, R., 2014. OpenFluor- an online spectral library of auto-fluorescence by organic compounds in the environment. Anal Methods 6 (3), 658–661. https://doi.org/10.1039/c3ay41935e.

[48] MacMillan, G.A., Clayden, M.G., Chetelat, J., Richardson, M.C., Ponton, D.E., Perron, T., Amyot, M., 2019. Environmental Drivers of Rare Earth Element Bioaccumulation in Freshwater Zooplankton. *Environ Sci Technol* 53 (3), 1650–1660. https://doi.org/10.1021/acs.est.8b05547.

[49] Pourmand, A., Dauphas, N., Ireland, T.J., 2012. A novel extraction chromatography and MC-ICP-MS technique for rapid analysis of REE, Sc and Y: Revising CI-chondrite and Post-Archean Australian Shale (PAAS) abundances. *Chem Geol* 291, 38–54. https://doi.org/10.1016/j.chemgeo.2011.08.011.

[50] Slack, J.F., Kelley, K.D., & Clark, J.L. (2004). *Whole Rock Geochemical Data For Altered And Mineralized Rocks, Red Dog Zn-Pb-Ag District, Western Brooks Range, Alaska*. (OPEN-FILE REPORT 2004-1372). Eastern Publications Group Retrieved from https://pubs.usgs.gov/of/2004/1372/2004-1372.html.

[51] Liu, H.Y., Guo, H.M., Xing, L.N., Zhan, Y.H., Li, F.L., Shao, J.L., Niu, H., Liang, X., Li, C.Q., 2016. Geochemical behaviors of rare earth elements in groundwater along a flow path in the North China Plain. J Asian Earth Sci 117, 33–51. https://doi.org/10.1016/j.jseae.2015.11.021.

[52] Hissler, C., Hostache, R., Ifffy, J.F., Pfister, L., Stille, P., 2015. Anthropogenic rare earth element fluxes into floodplains: Coupling between geochemical monitoring and hydrodynamic sediment transport modelling. *Comptes Rendus Geosci* 347 (5-6), 294–303. https://doi.org/10.1016/j.crte.2015.01.003.

[53] Bryan, S.E., Tipping, E., Hamilton-Taylor, J., 2002. Comparison of measured and modelled copper binding by natural organic matter in freshwaters. *Comp Biochem Physiol C-Toxicol Pharmacol* 133 (1-2), 37–49. https://doi.org/10.1016/S1532-0456(02)00083-2.

[54] Clune, J.W., Crawford, J.K., Chappell, W.T., Boyer, E.W., 2020. Differential effects of land use on nutrient concentrations in streams of Pennsylvania. Environ Res Commun 2 (11). https://doi.org/10.1088/2515-7620/abc97a.

[55] Shogren, A.J., Zarnetske, J.P., Abbott, B.W., Iannucci, F., Medvedeff, A., Cairns, S., Duda, M.J., Bowden, W.B., 2021. Arctic concentration-discharge relationships for dissolved organic carbon and nitrate vary with landscape and season. *Limnol Oceanogr* 66, S197–S215. https://doi.org/10.1002/lno.11682.

[56] Amon, R.M.W., Rinehart, A.J., Duan, S., Louchouarn, P., Prokushkin, A., Guggenberger, G., Bauch, D., Stedmon, C., Raymond, P.A., Holmes, R.M., McClelland, J.W., Peterson, B.J., Walker, S.A., Zhulidov, A.V., 2012. Dissolved organic matter sources in large Arctic rivers. *Geochim Et Cosmochim Acta* 94, 217–237. https://doi.org/10.1016/j.gca.2012.07.015.

[57] Lee, J.H., Byrne, R.H., 1993. Complexation of trivalent rare earth elements (Ce, Eu, Gd, Tb, Yb) by carbonate ions. *Geochim Et Cosmochim Acta* 57, 295–301.
[58] Millero, F.J., 1992. Stability-constants for the formation of rare-earth inorganic complexes as a function of ionic-strength. Geochim Et Cosmochim Acta 56 (8), 3123-3132. https://doi.org/10.1016/0016-7037(92)90293-R.

[59] Pokrovsky, O.S., Schott, J., 2002. Iron colloids/organic matter associated transport of major and trace elements in small boreal rivers and their estuaries (NW Russia). Chem Geol 190 (1-4), 141-179. https://doi.org/10.1016/S0009-2541(02)00115-8.

[60] Pourret, O., Dia, A., Davranche, M., Gruau, G., Hénin, O., Angée, M., 2007. Organo-colloidal control on major- and trace-element partitioning in shallow groundwaters: Confronting ultrafiltration and modelling. Appl Geochem 22 (8), 1568-1582. https://doi.org/10.1016/j.apgeochem.2007.03.022.

[61] Pokrovsky, O.S., Manasypov, R.M., Loiko, S.V., Krickov, I.A., Kopysov, S.G., Kolesnichenko, L.G., Vorobyev, S.N., Kirpotin, S.N., 2016. Trace element transport in western Siberian rivers across a permafrost gradient. Biogeosciences 13 (6), 1877-1900. https://doi.org/10.5194/bg-13-1877-2016.

[62] Liang, T., Li, K.X., Wang, L.Q., 2014. State of rare earth elements in different environmental components in mining areas of China. Environ Monit Assess 186 (3), 1499-1513. https://doi.org/10.1007/s10661-013-3469-8.

[63] Jiang, C.L., Li, Y.H., Li, C., Zheng, L.L., Zheng, L.G., 2022. Distribution, source and behavior of rare earth elements in surface water and sediments in a subtropical freshwater lake influenced by human activities. Environ Pollut 313. https://doi.org/10.1016/j.envpol.2022.120153.

[64] Sholkovitz, E.R., 1995. The aquatic chemistry of rare earth elements in rivers and Estuaries. Aquat Geochem 1 (1), 1-34. https://doi.org/10.1007/Bf01025229.

[65] Marginson, H., MacMillan, G.A., Grant, E., Gerin-Lajoie, J., Amyot, M., 2023. Rare earth element bioaccumulation and cerium anomalies in biota from the Eastern Canadian subarctic (Nunavik). Sci Total Environ 879, 163024. https://doi.org/10.1016/j.scitotenv.2023.163024.

[66] Vukov, O., Smith, D.S., McGeer, J.C., 2016. Acute dysprosium toxicity to *Daphnia pulex* and *Hyalella azteca* and development of the biotic ligand approach. Aquat Toxicol 170, 142-151. https://doi.org/10.1016/j.aquatox.2015.10.016.

[67] Techer, D., Grosjean, N., Sohm, B., Blaudez, D., Le Jean, M., 2020. Not merely noxious? Time-dependent hormesis and differential toxic effects systematically induced by rare earth elements in *Escherichia coli*. Environ Sci Pollut Res 27 (5), 5640-5649. https://doi.org/10.1007/s11356-019-07002-z.

[68] Yang, G., Wilkinson, K.J., 2018. Biouptake of a rare earth metal (Nd) by *Chlamydomonas reinhardtii* bioavailability of small organic complexes and role of hardness ions. Environ Pollut 243, 263-269. https://doi.org/10.1016/j.envpol.2018.08.066.

[69] Wei, Z.G., Yin, M., Zhang, X., Hong, F.S., Li, B., Tao, Y., Zhao, G.W., Yan, C.H., 2001. Rare earth elements in naturally grown fern *Dicranopteris linearis* in relation to their variation in soils in South-Jiangxi region (Southern China). Environ Pollut 114 (3), 345-355. https://doi.org/10.1016/S0269-7491(00)00240-2.

[70] Carroll, J.J., Lupsky, J.D., Mather, A.E., 1991. The solubility of carbon-dioxide in water at low-pressure. J Phys Chem Ref Data 20 (1), 1201-1209. https://doi.org/10.1063/1.555900.

[71] Hoffmann, L.J., Breitbarth, E., Boyd, P.W., Hunter, K.A., 2012. Influence of ocean warming and acidification on trace metal biogeochemistry. Mar Ecol Prog Ser 470, 191-205. https://doi.org/10.3354/meps10082.

[72] Takahashi, T., Sutherland, S.C., Sweeney, C., Poisson, A., Metzl, N., Tilbrook, B., Bates, N., Wanninkhof, R., Feely, R.A., Sabine, C., Olafsson, J., Nojiri, Y., 2002. Global sea-air CO flux based on climatological surface ocean CO, and seasonal biological and temperature effects. Deep-Sea Res Part II-Top Stud Oceanogr 49 (9-10), 1601-1622. https://doi.org/10.1016/S0967-0645(02)00003-6.

[73] Millero, F.J., Woosley, R., Ditrolio, B., Waters, J., 2009. Effect of ocean acidification on the speciation of metals in seawater. Oceanography 22 (4), 72-85. https://doi.org/10.5670/oceanogr.2009.98.

[74] Xu, Z.F., Han, G.L., 2009. Rare earth elements (REE) of dissolved and suspended loads in the Xijiang River, South China. Appl Geochem 24 (9), 1803-1816. https://doi.org/10.1016/j.apgeochem.2009.06.001.

[75] Goldstein, S.J., Jacobsen, S.B., 1988. Rare earth elements in river waters. Earth Planet Sci Lett 89 (1), 35-47. https://doi.org/10.1016/0012-821x(88)90031-3.

[76] Orhanovic, Z., Pokric, B., Furedi, H., Branica, M., 1966. Precipitation and hydrolysis of metallic ions. III. Studies on the solubility of yttrium and some rare earth hydroxides. Croat Chem Acta 38 (4), 269-276. (https://hrcak.srce.hr/file/305703).

[77] Lawrence, M.G., Kamber, B.S., 2006. The behaviour of the rare earth elements during estuarine mixing-revisited. Mar Chem 100 (1-2), 147-161. https://doi.org/10.1016/j.marchem.2005.11.007.

[78] Health Canada. (2009). Guidelines for Canadian Drinking Water Quality: Guideline Technical Document — Radiological Parameters. (Catalogue No. H128-1/10-614E- PDF). Ottawa, Ontario: Health Canada Retrieved from https://www.canada.ca/content/dam/canada/health-canada/migration/healthy-canadians/publications/healthy-living-vie-saine/water-radiological-radiologie-eau/alt/water-radiological-radiologique-eau-eng.pdf.

[79] Sarin, M.M., Krishnaswami, S., Somayajulu, B.L.K., Moore, W.S., 1990. Chemistry of uranium, thorium, and radium isotopes in the ganga-brahmaputra river system - weathering processes and fluxes to the Bay of Bengal. Geochim Et Cosmochim Acta 54 (5), 1387-1396. https://doi.org/10.1016/0016-7037(90)90163-F.

[80] Smedley, P.L., Kinniburgh, D.G., 2023. Uranium in natural waters and the environment: Distribution, speciation and impact. Appl Geochem 148. https://doi.org/10.1016/j.apgeochem.2022.105534.

[81] Chabaux, F., Bourdon, B., Riotte, J., 2008. Chapter 3: U-Series Geochemistry in Weathering Profiles, River Waters and Lakes. In: Krishnaswami, S., Cochran, J.K. (Eds.), Radioactivity in the Environment, Vol. 13. Elsevier, pp. 49-104. https://doi.org/10.1016/S1569-4860(07)00003-4.

[82] Health Canada. (2019). Guidelines for Canadian Drinking Water Quality: Guideline Technical Document — Uranium. (Catalogue No H144-13/12-2019E-PDF). Ottawa, Ontario: Health Canada Retrieved from http://www.canada.ca/en/health-canada/services/environmental-workplace-health/reports-publications/water-quality.html.

[83] Lafrance, I., Bandyayera, D., Charette, B., Bilodeau, C., & David, J. (2016). Géologie de la région du lac Brisson (SNRC 24A). (RG 2015-05). La Direction Générale de Géologie Québec, Gouvernement du Québec.

[84] Dupré, B., Gaillardet, J., Rousseau, D., Allegère, C.J., 1996. Major and trace elements of river-borne material: the Congo Basin. Geochim Et Cosmochim Acta 60 (8), 1301-1321. https://doi.org/10.1016/0016-7037(96)00043-9.

[85] Rabajczky, A., Namiesnik, J., 2014. Speciation of iron in the aquatic environment. Water Environ Res 86 (8), 741-758. https://doi.org/10.2175/106143014x13975035525906.

[86] Andersson, P.S., Dahlqvist, R., Ingri, J., Gustafsson, Ö., 2001. The isotopic composition of Nd in a boreal river: a reflection of selective weathering and colloidal transport. Geochim Et Cosmochim Acta 65 (4), 521-527. https://doi.org/10.1016/S0016-7037(00)00535-4.

[87] Caccia, V.G., Millero, F.J., 2007. Distribution of yttrium and rare earths in Florida Bay sediments. Mar Chem 104 (3-4), 171-185. https://doi.org/10.1016/j.marchem.2006.11.001.

[88] Sklyarova, O.A., Sklyarov, E.V., Och, L., Pastukhov, M.V., Zagarulko, N.A., 2017. Rare earth elements in tributaries of Lake Baikal (Siberia, Russia). Appl Geochem 82, 164-176. https://doi.org/10.1016/j.apgeochem.2017.04.018.

[89] Pourret, O., Davranche, M., Gruau, G., Dia, A., 2007. Organic complexation of rare earth elements in natural waters: Evaluating model calculations from ultrafiltration data. Geochim Et Cosmochim Acta 71 (11), 2718-2735. https://doi.org/10.1016/j.gca.2007.04.001.

[90] Catrouillet, C., Guenet, H., Pierson-Wickmann, A.C., Dia, A., LeCoz, M.B., Deville, S., Lenne, Q., Suko, Y., Davranche, M., 2020. Rare earth elements as tracers of active colloidal organic matter composition. Environ Chem 17 (2), 133-139. https://doi.org/10.1071/En19159.

[91] Tang, J.W., Johannesson, K.H., 2003. Speciation of rare earth elements in natural terrestrial waters: assessing the role of dissolved organic matter from the modeling approach. Geochim Et Cosmochim Acta 67 (13), 2321-2339. https://doi.org/10.1016/S0016-7037(02)01413-8.

[92] Ramos, S.J., Dinali, G.S., Oliveira, C., Martins, G.C., Moreira, C.G., Siqueira, J.O., Guilherme, L.R.G., 2016. Rare earth elements in the soil environment. Curr Pollut Rep 2 (1), 28-50. https://doi.org/10.1007/s40726-016-0026-4.

[93] Bro, R., 1997. PARAFAC. Tutorial and applications. Chemom Intell Lab Syst 38 (2), 149-171. https://doi.org/10.1016/S0169-7439(97)00032-4.

[94] Huguet, A., Vacher, L., Relexans, S., Saubusse, S., Froidefond, J.M., Parlanti, E., 2009. Properties of fluorescent dissolved organic matter in the Gironde Estuary. Org Geochem 40 (6), 706-719. https://doi.org/10.1016/j.orggeochem.2009.03.002.

[95] Shakil, S., Tank, S.E., Kokelj, S.V., Vonk, J.E., Zolkos, S., 2020. Particulate dominance of organic carbon mobilization from thaw slumps on the Peel Plateau, NT: Quantification and implications for stream systems and permafrost carbon release. Environ Res Lett 15 (11). https://doi.org/10.1088/1748-9326/abac36.

[96] Baken, S., Degryse, F., Verheyen, L., Merckx, R., Smolders, E., 2011. Metal complexation properties of freshwater dissolved organic matter are explained by its aromaticity and by anthropogenic ligands. Environ Sci Technol 45 (7), 2584-2590. https://doi.org/10.1021/es103532a.

[97] Lachaux, N., Otero-Fariña, A., Minguez, L., Sohm, B., Rétif, J., Châtel, A., Poirier, L., Devin, S., Pain-Devin, S., Gross, E., Giamberini, L., 2023. Fate, subcellular distribution and biological effects of rare earth elements in a freshwater bivalve under complex exposure. Sci Total Environ 905. https://doi.org/10.1016/j.scitotenv.2023.167302.

[98] Cory, R.M., McKnight, D.M., Chin, Y.P., Miller, P., Jaros, C.L., 2007. Chemical characteristics of fulvic acids from Arctic surface waters: Microbial contributions and photochemical transformations. J Geophys Res-Biogeosciences 112 (G4). https://doi.org/10.1029/2006jg000343.

[99] Möller, P., Dulski, P., De Lucia, M., 2021. REY patterns and their natural anomalies in waters and brines: the correlation of Gd and Y Anomalies. Hydrology 8 (3). https://doi.org/10.3390/hydrology8030116.

[100] Leybourne, M.I., Johannesson, K.H., 2008. Rare earth elements (REE) and yttrium in stream waters, stream sediments, and Fe-Mn oxyhydroxides: Fractionation, speciation, and controls over REE + Y patterns in the surface environment. Geochim Et Cosmochim Acta 72, 5962-5983.

[101] Pourret, O., Davranche, M., Gruau, G., Dia, A., 2008. New insights into cerium anomalies in organic-rich alkaline waters. Chem Geol 251 (1-4), 120-127. https://doi.org/10.1016/j.chemgeo.2008.03.002.

[102] Baud, A., Francus, P., Smol, J.P., Antoniades, D., Gregory-Eaves, I., 2023. Geochemical changes in Eastern Canadian lake sediment cores spanning the last ~150 years highlight a relative shift towards increased metals and erosive materials. Catena 225. https://doi.org/10.1016/j.catena.2023.107012.

[103] Piepenbrock, A., Kappler, A., 2013. Humic substances and extracellular electron transfer. In: Gescher, J., Kappler, A. (Eds.), Microbial Metal Respiration. Springer. https://doi.org/10.1007/978-3-642-32867-1_5.

[104] Sverjensky, D.A., 1984. Europium redox equilibria in aqueous-solution. Earth Planet Sci Lett 67 (1), 70-78. https://doi.org/10.1016/0012-821x(84)90039-6.
[106] Aubert, D., Stille, P., Probst, A., 2001. REE fractionation during granite weathering and removal by waters and suspended loads: Sr and Nd isotopic evidence. Geochim Et Cosmochim Acta 65 (3), 387–406. https://doi.org/10.1016/S0016-7037(00)00546-9.

[107] Vazquez-Ortega, A., Perdrial, J., Harpold, A., Zapata-Rios, X., Rasmussen, C., McIntosh, J., Schaap, M., Pelletier, J.D., Brooks, P.D., Amistadi, M.K., Chorover, J., 2015. Rare earth elements as reactive tracers of biogeochemical weathering in forested rhyolitic terrain. Chem Geol 391, 19–32. https://doi.org/10.1016/j.chemgeo.2014.10.016.

[108] MacMillan, G.A., Chetelat, J., Heath, J.P., Mickpegak, R., Amyot, M., 2017. Rare earth elements in freshwater, marine, and terrestrial ecosystems in the eastern Canadian Arctic. Environ Sci-Process Impacts 19 (10), 1336–1345. https://doi.org/10.1039/c7em00082k.

[109] Romero-Freire, A., Turlin, F., Andre-Mayer, A.S., Pelletier, M., Cayer, A., Giamberini, L., 2019. Biogeochemical Cycle of Lanthanides in a Light Rare Earth Element-Enriched Geological Area (Quebec, Canada). Minerals 9 (10). https://doi.org/10.3390/min9100573.

[110] Statistics Canada, 2023. Census Profile, 2021 Census of Population. Statistics Canada Catalogue, Ottawa.

[111] Piarulli, S., Hansen, B.H., Ciesielski, T., Zocher, A.L., Malzahn, A., Olsvik, P.A., Sonne, C., Nordtug, T., Jenssen, B.M., Booth, A.M., Farkas, J., 2021. Sources, distribution and effects of rare earth elements in the marine environment: current knowledge and research gaps. Environ Pollut 291. https://doi.org/10.1016/j.envpol.2021.118230.