# Tracing basin-scale circulation of Hg from multiple sources using paleolimnological, approaches, sediment long-term monitoring data and isotopic fingerprinting: a case study from Lake Maggiore (Italy/Switzerland)

Davide A.L. Vignati, Aldo Marchetto, Andrea Lami, Benedetta Ponti, Marina Camusso, Lucia Valsecchi, Angela Boggero, Simona Musazzi, Holger Hintelmann, Roberta Bettinetti, et al.

► To cite this version:

Davide A.L. Vignati, Aldo Marchetto, Andrea Lami, Benedetta Ponti, Marina Camusso, et al. Tracing basin-scale circulation of Hg from multiple sources using paleolimnological, approaches, sediment long-term monitoring data and isotopic fingerprinting: a case study from Lake Maggiore (Italy/Switzerland). 2025. hal-05360582

HAL Id: hal-05360582

https://hal.univ-lorraine.fr/hal-05360582v1

Preprint submitted on 12 Nov 2025

HAL is a multi-disciplinary open access
archive for the deposit and dissemination of sci-
entific research documents, whether they are pub-
lished or not. The documents may come from
teaching and research institutions in France or
abroad, or from public or private research centers.

L'archive ouverte pluridisciplinaire HAL, est
destinée au dépôt et à la diffusion de documents
scientifiques de niveau recherche, publiés ou non,
émanant des établissements d'enseignement et de
recherche français ou étrangers, des laboratoires
publics ou privés.

<figure><img src="image_1.png" /></figure>

Distributed under a Creative Commons Attribution 4.0 International License
1

2

3

4

5

6

7

8

9

10

11

12

13

14

15

16

17

18

19

20

21

22

23

24

25

26

27

28

29

30

31

32

33

34

35

36

37

38

39

40

41

42

43

44

45

46

47

48

49

50

Tracing basin-scale circulation of Hg from multiple sources using paleolimnological approaches, sediment long-term monitoring data and isotopic fingerprinting: a case study from Lake Maggiore (Italy/Switzerland)

Davide A.L. Vignati<sup>a*</sup>, Aldo Marchetto<sup>b</sup>, Andrea Lami<sup>b</sup>, Benedetta Ponti<sup>c</sup>, Marina Camusso<sup>d</sup>, Lucia Valsecchi<sup>d</sup>, Angela Boggero<sup>b</sup>, Simona Musazzi<sup>b</sup>, Holger Hintelmann<sup>e</sup>, Roberta Bettinetti<sup>c</sup>, Laura Marziali<sup>d*</sup>

<sup>a</sup> Université de Lorraine, CNRS, LIEC, F-57000 Metz, France

<sup>b</sup> National Research Council – Water Research Institute (CNR-IRSA), Largo Tonolli 50, 28922 Verbania, Italy

<sup>c</sup> Università degli studi Insubria, DISUIT, Dept. of human sciences and innovation for the territory, Via Valleggio 11, Como, Italy

<sup>d</sup> National Research Council – Water Research Institute (CNR-IRSA), Via del Mulino 19, 20861 Brugherio, Italy.

<sup>e</sup> Water Quality Centre, Trent University, Peterborough, K9J 0G2, ON, Canada

*Corresponding author: david-anselmo.vignati@univ-lorraine.fr

LIEC - UMR 7360

UFR SciFA - Université de Lorraine - Campus Bridoux - Bât. IBISE

8, rue du Général Delestraint

57070 Metz

France

*Corresponding author: laura.marziali@irsa.cnr.it

National Research Council – Water Research Institute (CNR-IRSA)

Via del Mulino 19

20861

Brugherio (MB)

Italy
Abstract

The ubiquitous nature of mercury contamination complicates the understanding of its behavior in aquatic ecosystems suffering from legacy pollution. By combining paleolimnological approaches, long-term monitoring and isotopic fingerprinting, we reconstructed the chronology of Hg contamination from multiple sources in sediment cores of Lake Maggiore (Italy/Switzerland). Anthropogenic Hg contamination spanned at least 400 years, from the early 17th century until the end of the 20th century, with inputs from artisanal and industrial gold mining, hat making and chlor-alkali production. Total Hg concentrations ranged from natural background levels of about 0.1 mg kg⁻¹ to 25 mg kg⁻¹ at the peak of Hg releases from a chlor-alkali plant in the first half of the 20th century. This industrial Hg had a distinctly lighter δ²⁰²Hg signature compared with natural Hg and moved as far as the southernmost part of the lake. Sediments impacted by gold mining activities had a δ²⁰²Hg signature indistinguishable from natural Hg, despite Hg concentrations much higher than background levels. Similarly, no specific δ²⁰²Hg signature could be determined for the Hg used in hat making and a combination of paleolimnological and monitoring information was used to establish the possible extent of Hg contamination from this source. Mercury contamination has decreased steadily since the early 2000s, but total Hg levels remain above background values, especially in the southern part of the lake. According to model calculations, Hg in sediments should return to background levels between 2040 and 2100, but the agreement between these estimates and recent observations varies across sampling site.

Keywords

Sediment contamination; Legacy mercury; chlor-alkali industry; gold mining; contamination chronology; sediment recovery

Highlights

- Evidence of Hg pollution starting from the 17th century (gold mining activities)

- Hg from an industrial point source is a major contributor to legacy Hg contamination

- Hg from an industrial point source has a distinct δ²⁰²Hg isotopic signature

- Combined lines-of-evidence allow to identify multiple Hg sources in sediments

- Chronology of Hg contamination allows estimates of time for full sediment recovery
1. Introduction

Mercury is a contaminant of global concern due to its capacity to undergo Long-Range Atmospheric Transport (LRAT) and to strongly bioaccumulate and biomagnify along food chains (Bank, 2020; Selin, 2009). Because of these undesirable properties, Hg is considered a priority pollutant under several regulatory frameworks, including the international binding treaty *The United Nations Minamata Convention on Mercury* – MCM. The MCM has come into force in 2017 to manage Hg-related environmental and health issues at a global scale (Bank, 2020). European legislation has identified mercury as a toxic, ubiquitous, persistent, and bioaccumulative substance (uPBTs) (Directive 2008/105/EC). The uPBT classification acknowledges that local or regional management actions may fail to ensure good chemical status for mercury in European water bodies because of persistent Hg inputs from global sources via LRAT. However, implementing global policies to reduce Hg emissions to the atmosphere, as well as releases into water and soils does not necessarily eliminate the needs for local measures. These considerations are pivotal where legacy Hg contamination from e.g., past industrial activities has left a clear footprint on an ecosystem (Covelli et al., 2017; Mao et al., 2023; Wiederhold et al., 2015). Furthermore, some studies reported a decoupling between the reduction in Hg emissions and the decrease in Hg in biota (Vijayaraghavan et al., 2014; Wang et al., 2019), suggesting that decreases in external Hg inputs are not necessarily mirrored in biota.

Although not informative of trends in biota, sedimentary Hg levels provide information on the contamination history of a given ecosystem and of possible environmental and health issues linked to persistent legacy Hg contamination (Bolaños-Alvarez et al., 2024; Covelli et al., 2017; Varekamp et al., 2003). Predicting the time required by sedimentary Hg levels to revert to natural background levels can support managers in deciding whether interventions are required to further reduce local Hg fluxes or to boost natural remediation processes (Lin et al., 2017). Equally important is the ability to distinguish natural Hg sources from local anthropogenic sources in controlling Hg dynamics in a given ecosystem; a feature that has been considerably advanced by the advent of multi-collector Inductively Coupled Plasma Mass Spectrometry (MC-ICP-MS) (Foucher et al., 2009; Schudel et al., 2018; Wiederhold et al., 2015).

Lake Maggiore is a southern subalpine lake, located across the border between Italy and Switzerland, that has been subject to industrial Hg pollution from a chlor-alkali plant over most of the 20th century (Grassi, 2012; Guilizzoni et al., 2012; Guzzella et al., 2018; Marziali et al., 2021a). During the early 1970s, maximum total Hg levels exceeded 30 mg kg⁻¹ dry weight for composite sediment samples collected in the most contaminated parts of the Lake (Baudo et al., 1981; Damiani and Thomas, 1974). A hat factory has also been in operation in the town of Ghiffa (Fig. 1) from 1881 to 1981. Although Hg was used to cure pelts in the hat making process during part of this period, the present paper will provide evidence in support of a limited impact from this source as compared with inputs from the chlor-alkali plant.

Following the raising awareness of the risks related to environmental pollution, a convention was signed by the Swiss Federal Authorities and the Italian Authorities for the creation of the Commission for the Protection of Italian Swiss Waters (CIPAIS) in 1972. The first CIPAIS report was published in 2002 (Guzzella et al., 2002) and documented the presence of high Hg concentrations in the central part of the lake. Mercury levels were also analysed in some fish species (Bertola and Bielli, 2002). Based on monitoring data obtained between 1999 and 2014, Guzzella et al. (2018) showed that the decrease of Hg levels in lacustrine sediments was not mirrored by a corresponding trend in fish muscles. Furthermore, the extent of sediment recovery from Hg pollution varied across sampling point in different parts of the lake (Guzzella et al., 2018). Contrary to lacustrine sediments, Hg levels in sediments collected from the main tributaries of the lake did not show clear decreasing trends between 2001 and 2018 (Marziali et al., 2021a). In several cases, both lacustrine and riverine sediments had Hg levels above the
stations, have further enhanced the understanding of the lake’s hydrodynamics and nutrient levels (Rogora et al., 2021).

<figure><img src="image_2.png" /><figcaption>Figure 1. Maps of Lake Maggiore catchment and basin with indication of known Hg point sources (chlor-alkali plant located in the town of Pieve Vergonte, mining area of Valle Anzasca, hat factory in the town of Ghiffa) and sampling points for the sediment cores. Sites 29 and 30B are located in Switzerland, while all the other sites are located in Italy. The checkered pattern in the inset corresponds to the area defined as a site of national interest (SIN) for priority remedial action by the Italian Ministry of Environment in January 2000. At the time of writing, consultations among parties are ongoing to narrow the extension of the SIN around the town Pieve Vergonte. No changes are foreseen as to the Lake areas currently included in the SIN (see text for more details).</figcaption></figure>
Lake Maggiore has a long history of pollution by both inorganic compounds such as Hg and persistent organic pollutants (POPs) including DDT and polychlorinated biphenyls (PCBs) (Guzzella et al., 2018). Industrial Hg discharges to Lake Maggiore likely began in the late 1910s when a chlor-alkali production line using Hg as a catalyst entered into service at a chemical plant located in the town of Pieve Vergonte (Fig. 1). In the mid-1930s, the chemical factory became an important hub for the production of chemicals with potential uses in chemical warfare (notably As₂O₃ and NaOH for the synthesis of As-based poisonous gases) and 68 Krebbs cells using Hg as a catalyst were installed. In 1954, the chlor-alkali production capacities were further developed to meet the growing requirements raised by various economic sectors after World War II. Between the 1960s and the mid-1970s, the Krebbs cells were progressively replaced by the more efficient De Nora cells. The overall productivity of the Pieve Vergonte factory started however to decline during the economic downturn of the 1970s. In 2004, the Italian government offered financial support to the owners to modernize the chlor-alkali production line from Hg-cells-based processes to mercury free membrane technology (Grassi, 2012; Grassi, 2013). However, the transition to the membrane technology was completed only in 2021 after the site was acquired by the Esseco Group that now runs the site via its division Altair Chemicals (https://stage.altairchimica.com/en/).

In 1996, the detection of DDTs levels above legal limits in fish from Lake Maggiore marked the beginning of regular monitoring programs on hazardous substances in the Lake Maggiore ecosystem by CIPAIS (Calderoni et al., 1999). The program was later expanded to include Hg and other pollutants. In January 2000, the Italian Ministry of the environment delimited an area comprising the industrial site in Pieve Vergonte, part of the River Toce and part of Lake Maggiore basin (Fig. 1) as a Site of National Interest for priority remedial action (Grassi, 2013; Ministerodell'Ambiente, 2000). At the time of writing, the Italian Ministry of Environment and Energy Security (MASE) is considering a revision of the areal extension of the SIN, but no changes are foreseen as to the area of Lake Maggiore presently included in the SIN perimeter.

Sediment cores for the present study were mostly collected from sites regularly monitored by CIPAIS and situated both inside and outside the area defined as SIN (Fig. 1). Sampling sites in the Borromean Gulf, directly receiving inputs of contaminated material from the River Toce, included one location for the reconstruction of historical Hg contamination (site ex13A) and three locations for evaluating the spatial and short-term temporal variability of Hg levels (sites 16, 17 and 51) (Fig. 1, Table S1). Other sites were located along the north to south axis in the main lake basin and were equally distributed among the northern (sites 29 and 30b), central (sites 54 and 55) and southern (sites 27 and 28) parts of the lake (Fig. 1 and Table S1). Sites ex13A, 16, 17, 51, 54, and 55 were inside the perimeter of the SIN (Fig. 1).

## 2.2. *Sampling and sample preparation*

Sediment cores were collected between 2006 and 2011 using a gravity corer operated from the research vessel of CNR-IRSA, transported back to the laboratory within 3 hours and stored at 4°C. Cores were opened longitudinally and immediately subsampled. In the remainder of the present paper, each core will be uniquely identified as explained in table S1.

Subsamples were obtained by slicing cores in either regular sections of 1 cm or in sections of variable thickness to account for different sedimentation rates at various locations. Regular intervals were used for cores sliced along their entire length to reconstruct the long-term evolution of Hg contamination, specifically cores LM06_ex13A, i.e., collected in 2006 at site ex13A, and LM06_27 (Table S1). For all the other cores (Table S1), sedimentation rates were established as described in section 2.3 and the thickness of sediment sections was then chosen to collect material covering the same span of time in all cores. The two approaches were combined for core LM11_28 (i.e. collected in 2011 at site 28), which was used for both long-term and synoptic study of Hg contamination.
## 2.3. Sediment chronology

Sediment chronology was established using different combinations of radiometric measurements ($^{137}$Cs, $^{210}$Pb, $^{14}$C), changes in diatom composition and abundance, and the presence of turbiditic deposits associated to an exceptional flood that occurred in the River Toce catchment in 2000.

For core LM06_ex13A, the age of sediment sections was established by lithostratigraphic correlation with another core collected nearby in the Borromean Gulf (45°54.84′N, 8°33.02′E) and dated using radiochemical methods ($^{137}$Cs, $^{210}$Pb, $^{14}$C) as described in Guilizzoni et al. (2012) and Putyrskaya and Klemt (2007).

For all other cores, the selected biological markers allowed to establish robust and coherent chronologies among cores. Indeed, the various locations show remarkable similarities in the composition of planktonic species despite differences as to the presence of benthic diatom species (Marchetto et al., 2004). The transition from small-sized *Stephanodiscus* populations to *Pantocsekiella comensis* (formerly *Cyclotella comensis*) dominated populations corresponds to the year 1989, while the reverse transition from *P. comensis* to *Stephanodiscus spp.* to the year 1963. Some of the cores collected in 2008, also contained turbiditic deposits related to the same major flood event in the River Toce (Kämpf et al., 2012). When such deposits could be clearly identified (Table S1), their top layer was taken as representative of the fall season in year 2000 and used to estimate the sedimentation rates between 2000 and 2008.

## 2.4. Mercury analysis

After core slicing, sediment samples were transferred to amber glass bottles, frozen and freeze-dried. The dried material was further homogenized in an agate or zircon mortar and stored in a desiccator pending analysis. Total mercury concentrations were determined according to US EPA method 7473 (USEPA, 1998) using an automated Hg analyzer (AMA254, FKV, Bergamo, Italy). Analytical accuracy was checked with reference materials BCR280 (0.670 ± 0.019 mg kg⁻¹), GBW07305 (0.1 mg kg⁻¹), and an internal reference sediment labelled S7 (1.72 ± 0.04 mg kg⁻¹) (Guilizzoni et al., 2012). Replicate measurements were performed on selected sections and at least once for each analytical run.

Samples for analysis of Hg isotope ratios were digested using concentrated HNO₃/H₂SO₄ (7:3 v/v). Between 20 and 200 mg of dried sample material was digested with 5 mL of the acid mixture in open glass vessels on a hot plate at 120 °C for 6–8 h. After cooling, the samples were appropriately diluted with MilliQ water to reach final Hg concentration of 10 ng mL⁻¹ for optimum performance and accuracy of the Hg isotope ratio measurements.

Hg isotope ratios in the digested solutions were measured by multi-collector inductively coupled plasma mass spectrometry (MC-ICPMS, Neptune). The mass bias of MC-ICPMS was corrected by both the internal NIST 997 Tl standard using an exponential fractionation law and the NIST 3133 Hg standard-sample bracketing method. The bracketed NIST 3133 solutions were matched to the digested solutions within 10% in both matrix and Hg concentrations. Hg isotope ratio is expressed in delta notation (δ, ‰) by referring to the bracketed NIST 3133 Hg standard (Blum and Bergquist, 2007):

$$ \delta^{XXX} \text{Hg (‰)} = \left( \left( \frac{^{XXX} \text{Hg}}{^{198} \text{Hg}} \right)_{\text{Sample}} - \left( \frac{^{XXX} \text{Hg}}{^{198} \text{Hg}} \right)_{\text{NIST 3133}} - 1 \right) \times 1000 \quad (1) $$

where XXX is the Hg isotope mass between 198 and 202, (XXXHg/¹⁹⁸Hg)sample is the measured isotope ratio of the sample, and (XXXHg/¹⁹⁸Hg)NIST3133 is the average of isotope ratios of bracketed NIST3133 solutions measured before and after each sample. Mass independent fractionation (MIF) is indicated by capital delta (Δ, ‰) notation, which is the difference
between measured values of δ¹⁹⁹Hg, δ²⁰⁰Hg, δ²⁰¹Hg and those predicted from δ²⁰²Hg using the kinetic Mass Dependent Fraction (MDF) laws (Blum and Bergquist, 2007):

$$ \Delta^{xxx} \text{Hg} = \delta^{xxx} \text{Hg} - \beta_{xxx} \times \delta^{202} \text{Hg} \quad (2) $$

The mass dependent scaling factor βxxx is 0.252, 0.502 and 0.752 for ¹⁹⁹Hg, ²⁰⁰Hg and ²⁰¹Hg, respectively.

## 2.5. Quality control procedures

Average accuracy of total Hg measurements was 111.6 ± 9.2 % (range 94.8–121.9 %, n = 10) for S7, 101.8 ± 8.5 % (range 89.7 – 111.9 %, n = 8) for BCR680 and 97.7 ± 5.2 % (range 87.9 – 106.4 %, n = 8) for GBW. Replicate analyses for S7, GBW had a relative standard deviation below 5% (Table S2). Analytical precision (duplicate or triplicate analyses) for sediment samples was usually below 10%, with the highest values likely originating from residual sample heterogeneity (Tables S3–S6).

The secondary standard UM-Almaden solution gave average Hg isotope ratios of -0.59 ± 0.04‰, -0.03 ± 0.04‰, 0.04 ± 0.04‰ (2SD, n = 8) for δ²⁰²Hg, Δ¹⁹⁹Hg and Δ²⁰⁰Hg, respectively, which were consistent with the previously reported values (Blum and Bergquist, 2007; Sun et al., 2013).

## 2.6. Data analysis and statistics

In cores used for the long-term reconstruction of Hg contamination (LM06_ex13a, LM06_27, and LM11_28), temporal trends in the decrease of Hg concentrations were evaluated by fitting hyperbolic, 2-parameter models:

$$ y = \frac{a \cdot b}{(b+x)} \quad (3) $$

or single exponential decay, 2-parameter models:

$$ y = a \cdot e^{-b \cdot x} \quad (4) $$

where x is the time in years as established according to sediment chronology for a given sediment section, y is the Hg concentration in the corresponding sediment section, and ‘a’ and ‘b’ are fitted model parameters.

For each core, models were fitted to Hg concentrations measured in sections covering the span from the most recent peak of contamination to the sediment surface at the time of sampling (see section 3.1.4, 3.2.3, and 3.2.5 for details). The dynamic fitting option in SigmaPlot version 14 was used for model fitting.

In cores sampled for synoptic studies of Hg contamination, differences in Hg concentrations between 2008 and 2011 were checked for each individual site using the Student t-test after verification for the homogeneity of variances (F test). Following the presence of non-normally distributed data (Shapiro-Wilk test), differences across sites were verified using one-way Anova by ranks followed by pairwise comparison with the Student-Newman-Keuls (SNK) and Tukey tests. The reasons for using two post-hoc comparison methods will be explained in section 3.3. Statistical analysis was performed using SigmaPlot ver 14.
364 **3. Results and discussion**

365 **3.1. Geochronology of Hg contamination in the Borromean Gulf (core LM06_ex13A)**

366 **3.1.1. Sediment chronology**

367 The chronology of core LM06_ex13A has been previously established by Guilizzoni et al. (2012). Briefly, radiometric measurements showed two peaks of ¹³⁷Cs at 11.5 cm and 24.5 cm, corresponding to the accident at the nuclear power plant in Chernobyl in 1986 and to post world-war II nuclear tests in 1963, respectively (Fig. 2). A ¹⁴C measurement dated section 66–69 cm at 1620 ± 70 AD (Fig. 2). A sedimentation rate of 0.53 cm yr⁻¹ (Table S1) was established for recent sediments based on the 1963 marker. The corresponding value for pre-1620 sediments was calculated at 0.18 cm yr⁻¹. Extrapolating the latter rate to the bottom of the core resulted in an estimated calendar age of circa 1400 AD. Although Hg contamination in this core was already examined by Guilizzoni et al. (2012), the present study will provide a more detailed understanding on the full chronology of Hg levels from the pre-industrial period until the recovery phase following massive anthropogenic inputs during the 20th century.

378

379 **3.1.2. Background mercury levels in the Borromean Gulf**

380 Mercury levels were around 0.05–0.1 mg kg⁻¹ in the four oldest sections of core LM06_ex13A (89–90, 94–95, 99–100 and 106–107 cm) and their average Hg level (arithmetic mean ± one standard deviation) was 0.044 ± 0.022 mg kg⁻¹ (Table S3, Fig. 2). These values agree with those proposed by Viel and Damiani (1985) based on 126 surface sediment samples collected throughout Lake Maggiore and with those measured in cores collected in the northernmost part of the lake far from known anthropogenic sources of Hg (see section 3.3). Given the age of the four bottom sections of the core (circa 1400–1500 AD), the corresponding Hg values clearly predate industrial influences. However, artisanal gold mining is believed to have occurred in Anzasca Valley (River Toce catchment) since Roman times, although historical evidence exists only starting from the XII–XIII century (Agostinetti et al., 1995; Del Soldato, 1986; Del Soldato, 1989).

391 The values measured in the lower part of the core are compatible with the presence of igneous rocks in the upper part of the River Toce catchment (Barbanti and Ambrosetti, 1989).
392 Mercury levels in igneous rocks range from 0.004 mg kg⁻¹ for ultramafic rocks to 0.08 mg kg⁻¹ for granites, while Hg content in sedimentary rocks varies between 0.16 for limestones and 0.29 for sandstones (Table S7). Earlier investigations (Baudo et al., 1981; Muntau, 1981) mention the presence Hg-rich rocks as a possible cause of high geogenic levels of mercury in the River Toce catchment. However, the average Hg content of all rock types present in the River Toce catchment is 0.24 ± 0.37 mg kg⁻¹ (arithmetic mean ± 1 s.d.) with a median of 0.08 mg kg⁻¹, a 1st quartile of 0.06 mg kg⁻¹ and a 3rd quartile of 0.22 mg kg⁻¹ (Table S7). Paragneiss (1.30 mg Hg kg⁻¹) and serpentine-amphibolite (1.5 mg Hg kg⁻¹) do occur in the zone of Gignese and in the Anzasca Valley, respectively. However, mercury levels in sediments collected in 2017 in the River Anza, draining the Anzasca Valley, were lower than those in the River Toce upstream of its confluence with the River Anza: 0.017 ± 0.002 mg kg⁻¹; in River Anza vs. 0.034 ± 0.004 mg kg⁻¹ in the River Toce (our unpublished data). Based on the available information, background Hg levels in the sediments of the Borromean Gulf are likely below 0.1 mg kg⁻¹. Concentrations above this value in recent sediments may be considered as indicative of persistent legacy contamination from anthropogenic sources.

393
394
395
396
397
398
399
400
401
402
403
404
405
406
407
408
409
<figure><img src="image_3.png" /></figure>

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

<figure><img src="image_5.png" /></figure>

Figure 2. Panel A. Chronology of Hg contamination in core LM06_ex13A. Dates next to the left axis correspond to known markers of calendar age (in blue) or estimated dates for peak Hg contamination (in orange). Different anthropogenic activities are indicated as the possible causes of increased Hg levels in the various sections of the core. Sediment sections post-1970 are considered representative of the recovery period of Lake Maggiore ecosystem considering the progressive decommissioning of the chlor-alkali plant in the River Toce catchment. Panel B. Graphical representation of the hyperbolic decay model of Hg concentrations between 1970 and 2006 AD (see equation 5 in the text), including measured Hg levels (black dots), fitted model (continuous black line), 95% confidence intervals of the model equation (blue lines), and 95% prediction intervals (red lines). Panel C. Simulation of the decrease of Hg concentrations according to equation 5 from 1972 (latest peak of Hg contamination) to the estimated Hg background levels of 0.1 mg kg⁻¹.

The proposed Hg background value of 0.1 mg kg⁻¹ is markedly lower than the level of 0.91 ± 0.21 mg kg⁻¹ proposed by Baudo et al. (1981) based on sediments collected in Lake Mergozzo in 1975. However, Locht and Muntau (1981) warned against using sediments from Lake Mergozzo as an unpolluted reference for Lake Maggiore because of their higher content in organic matter. Furthermore, unpublished data obtained by our group from a core collected in Lake Mergozzo in 2011 suggest that it also suffered from Hg contamination, although an exact chronology was not established for this core (Table S8).

### 3.1.3. Detailed Chronology of anthropogenic Hg contamination

Mercury concentration started to increase above background levels from sections 74–75 cm (0.638 ± 0.08 mg kg⁻¹) and 69–70 cm (0.493 ± 0.004 mg kg⁻¹), the latter corresponding to circa 1620 AD (Fig. 2). This increase remained visible up to 49–50 cm (about 1 mg kg⁻¹) and, according to historical reports could have originated from Hg use in artisanal gold extraction in several valleys close to Mount Rosa (River Toce catchment) (Giuliani et al., 2015; Madau et al., 1986). The increase became more marked from 44–45 cm following the onset of industrial gold mining extraction that lasted until 1961 (Del Soldato, 1986; Del Soldato, 1989). Because industrial releases of Hg from the Pieve Vergonte chlor-alkali plant started at the beginning of the 20th century, we attempted to estimate the relative contribution of gold mining vs. chlor-alkali production to the total Hg load to the Borromean Gulf.

One of the most active gold mines was in Pestarena (Macugnaga, Anzasca Valley), where Hg was used for gold extraction until 1945, when it was replaced by cyanide. This mine
produced about 1,500 kg of gold between 1938 and 1945 AD for an estimated Hg release between 75 and 1,500 kg (Camusso and Borsani, 2004). The lower figure assumes a release of 5% of the used Hg per kilogram of extracted gold, while the upper figure represents a worst-case scenario of 100% Hg release (Lacerda, 1997). For the worst-case scenario, the mercury content in sediments deposited over the period 1938–1945 should be about 0.860 mg kg⁻¹, assuming a thickness of 4 cm for the sediment layer deposited during this period (sedimentation rate of 0.57 cm yr⁻¹, Table S1), a sediment density of 2.5 kg m⁻³, and a homogeneous distribution of Hg across the Borromean Gulf using a surface area of 17.5 km² (Table S9). Other mines likely contributed additional Hg loads to the Borromean Gulf during the same period. However, about 62,500 kg of Hg (i.e., a production of at least 62,500 kg of gold) would have been necessary to raise Hg sediment levels to the values of 25 mg kg⁻¹ measured in the 32–33 cm section (circa 1950) at the peak of chlor-alkali production. The estimated Hg content of 0.860 mg kg⁻¹ is also about 4-fold lower than the concentration measured in section 39.5 cm (Table S3) that can be dated at around 1937 AD (no Hg measurements available for section 38.5 cm; Table S3). It is therefore highly likely that Hg levels exceeding 1 mg kg⁻¹ in sediments deposited during the 20th century were mostly caused by discharges from the chlor-alkali plant in Pieve Vergonte.

The absolute Hg maximum observed at 32–33 cm (circa 1950) was followed by a relative maximum during the 1970s (Fig. 2). The relative maximum (about 16 mg kg⁻¹) agreed well with Hg levels between 13.4 and 32.8 mg kg⁻¹ reported for surface sediments collected in the Borromean Gulf between 1961–1962 and 1986 (Baudo et al., 1981; Damiani and Thomas, 1974; Rossi et al., 1993). These previous studies used homogenized sediment samples integrating the upper 10 cm of sediments and were unable to highlight the temporal variability in Hg levels between the 1950s and the 1970s (Fig. 2). The observed fluctuations could reflect periods of higher and lower mineral inputs from River Toce due to hydrological factors as well as changes in the activity levels and operation of the chlor-alkali plant (Grassi, 2012). Starting from the 1970s, Hg levels decreased steadily until the beginning of the 1990s (8–10 cm) and then stabilized around 0.4 mg kg⁻¹; i.e., about 4- to 8-fold above the background levels. According to Grassi (2012), production processes using Hg as a catalyst were still in use as of 2006 and Hg atmospheric emissions were reported until 2017 (Fantozzi et al., 2021). In 2021, the Hg-catalytic cells were eventually replaced by a membrane-based chlor-alkali production process.

### 3.1.4. Modelling Hg concentration decrease

The decrease in Hg concentrations during the period 1972–2006 was best described by a hyperbolic-type decay:

$$y = \frac{(-0.0069 \pm 0.0003) \times (-1970.95 \pm 0.0425)}{(-1970.95 \pm 0.0425) + x} \quad (5)$$

where x is the calendar year and y is the Hg concentration in mg kg⁻¹. All model coefficients were highly significant (P < 0.0001), the adjusted R-square was 0.99 and the standard error of the estimate 0.2263, providing an overall superior fit compared to the alternative exponential decay model (Table S10). The latter will not be further discussed. According to the hyperbolic decay pattern, the background mercury level of 0.1 mg kg⁻¹ would be eventually reached in the year 2107 AD (Table S10).

Subsequent monitoring studies coordinated by the CIPAIS between 2014 and 2023 (Marziali et al., 2024; Marziali et al., 2015; Marziali and Valsecchi, 2022; Marziali et al., 2018) provide information on the more recent evolution of Hg concentration in the Borromean Gulf (Table S11). In particular, site 13 had coordinates very similar to site LM06ex_13A (see Table S1 and Table S11). Mercury concentrations in the topmost sediment layer of cores collected at site 13
496

(Table S11) therefore provide a good basis for assessing the predictive ability of equation 5.
497
Measured vs. modelled Hg concentrations (in mg kg⁻¹) were 0.378 vs. 0.316 for 2014, 0.225 vs.
498
0.295 for 2017, 0.183 vs. 0.271 for 2021 and 0.150 vs. 0.261 for 2023. While both experimental
499
observations and model calculations confirm decreasing Hg contamination, experimental data
500
suggest that the actual rate of recovery in the Borromean Gulf may be faster than predicted
501
using historical data. This discrepancy likely arises from the interruption of Hg emissions in
502
2017 and calls for more spatially refined models to understand the future evolution of Hg
503
concentrations in the Borromean Gulf.
504

505

## 3.2. Geochronology of Hg contamination in the southern part of Lake Maggiore (cores LM06_27 and LM11_28)

506

507

### 3.2.1. Sediment chronology

508

In core LM06_27, the transitions from *Stephanodiscus* to *Pantocsekiella* and from *Pantocsekiella* to *Stephanodiscus* were observed at 3.5 (1989 AD) and 7.5 cm (1963 AD), respectively. Because of the low sedimentation rate, only the 1963 AD marker was considered to establish the sediment chronology. A constant sedimentation rate of 0.16 cm yr⁻¹ (Table S1) was assumed for the entire core, dating the lowermost section at 1434 AD.

509

510

511

512

513

In core LM11_28, only the *Stephanodiscus* to *Pantocsekiella* transition (1989 AD) was observable at 5 cm. A sedimentation rate of 0.23 cm yr⁻¹ was used throughout the core length (Table S1), resulting in a relatively young age of *circa* 1820 AD at the bottom (section 42.5–42.5 cm).

514

515

516

517

We acknowledge that the assumption of constant sedimentation rate may not be valid, especially for sediments that may predate gold-mining activities in the River Toce catchment (see section 3.1) and other changes in land use within the Lake Maggiore catchment. Extrapolating sedimentation rates along the whole core length results in some differences in the estimated age of Hg peak contamination among core LM06_ex13A from the Borromean Gulf vs. cores LM06_27 and LM11_28 collected in the southern part of Lake Maggiore (Figs. 2, 3, 4). These chronological discrepancies can be both a methodological artefact due to sedimentation rates actually varying over time and a consequence of the time required for Hg inputs from the River Toce to reach this part of the lake.

518

519

520

521

522

523

524

525

526

527

### 3.2.2 Chronology of anthropogenic Hg contamination in core LM06_27

528

529

Mercury levels remained indistinguishable from the proposed background of < 0.1 mg kg⁻¹ up to section 49–50 cm that was dated *circa* 1700 AD (Fig 3, Table S1). From this depth, Hg concentrations continuously increased until 1784 AD (35–36 cm) and then fluctuated between a minimum of approximately 0.5 mg kg⁻¹ to a maximum of 3.35 mg kg⁻¹ until the beginning of the 20th century (16–17 cm) (Fig 3). These levels are comparable with those measured in core LM06_ex13A during periods of known artisanal and industrial gold mining in the River Toce basin (Fig. 2), although Hg concentrations in core LM06_27 were more variable than in core LM06_ex13A. Some fluctuations in Hg levels could have been overlooked in the chronology of core LM06_ex13A due to the lower resolution of the Hg profile during the gold mining period (Fig. 2 and section 3.1).

530

531

532

533

534

535

536

537

538

539

Two distinct peaks in Hg concentration occurred during the 20th century at 11.5 cm (17.3 mg kg⁻¹) and 8.5 cm (13.9 mg kg⁻¹), followed by a steady decrease up to the core surface. Despite some differences in the estimated age of Hg peak contamination (Figs. 2 and 3), the similar patterns observed in cores LM06_27 and LM06_ex13A suggest that Hg inputs from River Toce were transported outside the Borromean Gulf toward the southern part of Lake Maggiore. Results of MC-ICP-MS analyses (section 3.4) will provide further support to this hypothesis.

540

541

542

543

544

545

The possibility of Hg being transported outside the Borromean Gulf also agrees with known water transport patterns. In Lake Maggiore, water globally moves from north to south and from
west to east (i.e., from the Borromean Gulf to the main lake basin) down to depths of 130 m and 120 m, respectively (Ambrosetti et al., 2009; Baudo et al., 1981). Dissolved and colloidal Hg inputs can therefore move out of the Borromean Gulf thanks to the advective transport of water masses. The same trend may hold for Hg bound to suspended particulate matter (SPM).

According to Viel and Damiani (1985), sedimentary deposits in the southern part of Lake Maggiore are dominated by silts with a mode of 10–6 µm. Considering a depth of 85 m at the sampling site (Table S1) and a water temperature of 5 °C, calculations based on the Stokes’ law estimate that a quartz particle ($ρ_s$ = 2600 kg m⁻³) with a 10 µm diameter would take 17 days to reach the sediment surface, while a less dense algal particle ($ρ_s$ = 1200 kg m⁻³) will need 137 days (Table S12). These figures increase to 48 and 380 days for particles with a diameter of 6 µm, although higher water temperatures would decrease water viscosity and result in faster sedimentation. Despite the limitations and assumptions of Stokes’ law (e.g., no consideration of turbulent transport and water stratification, assumption of spherical particles), the overall water circulation pattern strongly supports the hypothesis that fine silts and colloids are transported from the River Toce to the southern part of the lake. Viel and Damiani (1985) also report that sediments in the southern part of the lake are dominated by autochthonous material with a large proportion of diatom frustules, although cores show an alternance of autochthonous and allochthonous material. Considering the importance of zooplankton in the Hg cycle (Marziali et al., 2022), incorporation of Hg in less dense biological particles may have facilitated the southward transport of anthropogenic Hg.

<figure><img src="image_6.png" /><figcaption>Figure 3. Panel A: Chronology of Hg contamination in core LM06_27. Panel B: Graphical representation of the hyperbolic decay model of Hg concentrations between 1953 and 2006 AD (equation 6) including measured Hg levels (black dots), fitted model (continuous black line), 95% confidence intervals of the model equation (blue lines), and 95% prediction intervals (red lines). Panel C: Simulation of the decrease of Hg concentrations from 1953 to the estimated Hg background level of 0.1 mg kg⁻¹ according to equation 6.</figcaption></figure>
### 3.2.3. Modelling Hg concentration decrease in core LM06_27

Both hyperbolic and exponential decay models could effectively describe the temporal decline in Hg concentration in core LM06_27 between 1953 and 2006 AD, although the hyperbolic model was slightly better in terms of adjusted R square (Table S10).

The hyperbolic decay of Hg concentrations (equation 6) would further support the hypothesis that inputs from River Toce have been the main source of Hg to this site. However, Hg levels estimated using equation 6 for the year 2011 (0.867 mg kg⁻¹) were twice than those measured in a core collected at the same site in 2011: 0.461 mg kg⁻¹ for the top 1 cm section (see section 3.3). According to equation 6, it would take until 2485 AD to reach background-like Hg concentrations in this part of the lake; a time much longer than estimated for core LM06_ex13a (Table S10). This longer recovery time can reflect the lower sedimentation rate at this site compared with the Borromean Gulf (Table S1), resulting in a slower deposition of clean material on the top of contaminated sediments.

Other factors contributing to longer recovery times in core LM06_27 include the possible presence of additional Hg sources close to the sampling site and sediment focussing due to lower water depths in the southern part of the basin (Table S1). Core LM06_27 may have received Hg inputs from the Bardello tributary, which originates from the heavily anthropized Lake Varese (Baudo et al., 1992; Guilizzoni and Lami, 1986). Sediment samples from the Bardello River have Hg concentrations comparable to those of the river Toce for the fraction < 63 µm (Marziali et al., 2021a). However, the average flow of Bardello river is one order of magnitude lower than that of the River Toce (Ciampittiello, 1999; Marziali et al., 2021a). Consequently, inputs from the river Bardello are likely to have a limited influence on the overall Hg budget in the southern part of Lake Maggiore.

$$y = \frac{(-0.0275 \pm 0.0031) \times (-1949.17 \pm 0.4839)}{(-1949.17 \pm 0.4839) + x} \quad (6)$$

Alternatively, it is instructive to examine the possibility of Hg concentration decreasing according to an exponential decay law (equation 7):

$$y = 6.56^{44} \times e^{(-0.0516 \pm 0.015)x} \quad (7)$$

Despite a slightly lower R² value for this model (Table S10), equation 7 gives a theoretical Hg concentration of 0.564 mg kg⁻¹ for the year 2011 and predicts that Hg levels should reach background values as early as 2044 AD. The concentration estimate is in better agreement with the experimental value of 0.461 mg kg⁻¹ measured in 2011 (see section 3.3) and with the estimates for core LM06_ex13A for the time to fully recover from Hg contamination (Table S10). Unfortunately, monitoring in this area has been discontinued after the 2011 campaign in favour of more detailed follow up studies in the Borromean Gulf.

### 3.2.4. Chronology of anthropogenic Hg contamination in core LM11_28

The age for the deepest section of this core was estimated at 1824 AD, which is much more recent than the bottom sections of cores LM06_ex13A and LM06_27. However, background-like Hg levels (< 0.1 mg kg⁻¹) were measured up to 31 cm (1876 AD). Moving upwards, mercury levels regularly increased to a maximum of 13.8 mg kg⁻¹ at 14 cm (1950 AD). A second, relative maximum of 8.6 mg kg⁻¹ was recorded at 9 cm (1972 AD). The estimated dates for peak Hg levels are in excellent (fair) agreement with the patterns observed in cores LM06_ex13A (LM06_27). Again, the available data support the hypothesis of Hg transport
from the Borromean Gulf to the southern part of the lake, especially during the period of maximum activity of the chlor-alkali plant.

As already observed for core LM06_27, Hg levels in surface sediments of core LM11_28 were higher than in core LM06_ex13A collected from the Borromean Gulf; i.e. closer to (legacy) Hg anthropogenic sources. Local tributaries are unlikely to contribute substantial Hg loads at site LM28. Besides the river Bardello (section 3.2.3), other local tributaries are rivers Erno, for which we could not find information on Hg levels, and Vevera. Sediments collected in river Vevera had Hg levels of 0.264 mg kg⁻¹ in February 1998 and 0.119 mg kg⁻¹ in September 1998 for the fraction < 63 µm (Menaballi, 1999). By comparison, Hg level in core LM28_11 at the depth of 3 cm (corresponding to circa 1998 AD) were 1.97 mg kg⁻¹, confirming a limited role of local tributaries in determining the high Hg levels in this part of the lake. Sediment focussing or enhanced sediment mixing likely contribute to the persistence of high Hg levels at site LM28 because of the shallow depth in this part of the lake (Table S1).

<figure><img src="image_7.png" /><figcaption>Figure 4. Panel A: Chronology of Hg contamination in core LM11_28. Panel B: Graphical representation of the hyperbolic decay model of Hg concentrations between 1972 and 2011 AD (equation 8) including measured Hg levels (black dots), fitted model (continuous black line), 95% confidence intervals of the model equation (blue lines), and 95% prediction intervals (red lines). Panel C: Simulation of the decrease of Hg concentrations from 1972 to the estimated Hg background levels of 0.1 mg kg⁻¹ according to equation 8.</figcaption></figure>

### 3.2.5. Modelling Hg concentration decrease in core LM06_28

The decrease in mercury concentrations from section 8.5–9.5 cm (1972 AD) to the surface of core LM28_11 was best described by an exponential decay model (Table S10; Fig. 4b).

$$y = 9.61^{44} \times e^{(-0.0515 \pm 0.0038)x} \quad (8)$$

According to equation 8, Hg concentrations should have decreased from approx. 1 mg kg⁻¹ (2011 value) to 0.866, 0.742, 0.604, and 0.544 mg kg⁻¹ in 2014, 2017, 2021, and 2023 AD,
respectively. These estimates compare favourably with experimental data from cores collected at the same site in between 2014 and 2023 despite differences in the thickness of sediment sections used for Hg analysis (Table 1, Table S11). The same model predicts that Hg levels at this site should drop to 0.1 mg kg⁻¹ by 2056 AD (Fig 4c). However, model results seem to slightly underestimate the speed of recovery compared with experimental data, similarly to what observed for core LM06_ex13A. These observations indicate that further monitoring remains desirable to follow changes in Hg levels in what is now the most contaminated part of the lake basin (see also section 3.3).

The alternate hyperbolic decay model (see Table S10) provided an overall good fit to the data, but predicted a recovery time that seems unrealistically long given the trends observed at sampling site LM28 (Fig. 4).

<table><caption>Table 1. Comparison of modelled (Hg mod, equation 8) vs. measured Hg concentrations in cores collected at site LM28 during 2011–2023.</caption><thead><tr><th>Sampling (year)</th><th>Hg<sub>mod</sub> (mg kg<sup>-1</sup>)</th><th>Hg<sub>obs</sub> (mg kg<sup>-1</sup>)</th><th>Sediment thickness (cm)</th><th>Time interval</th><th>Reference</th></tr></thead><tbody><tr><td>2011</td><td>1.01</td><td>1.1</td><td>0.5</td><td>2008–2011</td><td>1</td></tr><tr><td>2014</td><td>0.866</td><td>1.0</td><td>1.9</td><td>2004–2014</td><td>2</td></tr><tr><td>2017</td><td>0.742</td><td>0.71</td><td>0–0.35</td><td>2016–2017</td><td>3</td></tr><tr><td>2021</td><td>0.604</td><td>0.547</td><td>1</td><td>2018–2021</td><td>4</td></tr><tr><td>2023</td><td>0.544</td><td>0.483</td><td>1</td><td>2020–2023</td><td>5</td></tr></tbody></table>

1, Present study; 2 (Marziali et al., 2015), 3 (Marziali et al., 2018), 4 (Marziali and Valsecchi, 2022), 5 (Marziali et al., 2024). Raw data for cores collected in 2014–2023 can be found in Marziali et al. (2025).

### 3.3. Synopsis of Hg contamination in Lake Maggiore sediments

The series of cores collected in 2008 and 2011 (Table S1) provided comparable snapshots of Hg levels in the whole Lake Maggiore basin (Fig. 5; Table S6). At each location, average Hg concentrations for the three upper sediment layers did not show statistically significant differences among the two sampling periods (P > 0.05, Student-t test). The following discussion will therefore focus on differences across sampling sites based on the combined results from both campaigns. For the remainder of this section, Hg concentration will be given as the arithmetic means ± one standard deviation for each site with n = 6 at all cases (Table S13).

Mercury concentrations were significantly different across sites (P < 0.001, Kruskal-Wallis Anova by ranks). However, results for pairwise post-hoc comparisons between sites were sensitive to the chosen method, with the Student-Newman-Keuls (SNK) test being less conservative than the Tukey test (Table S14). In our opinion, the SNK test better captured the actual differences among sites and its results will be used in the following discussion.

The lowest Hg levels were found in cores 29 and 30b collected in the northern part of the lake at locations far from known anthropogenic Hg sources (Fig. 1). Mercury levels in core 29 (0.032 ± 0.011 mg kg⁻¹) were comparable to i) pre-industrial background levels in lake sediments (see section 3.1) and ii) concentrations in the fine sediment fraction (< 63 µm) of the Ticino and Maggia tributaries located in the Swiss part of Lake Maggiore basin (Marziali et al., 2021a; Menaballi, 1999). Mercury levels in cores 30b (0.089 ± 0.013 mg kg⁻¹) were significantly higher than at site 29 (P = 0.004) despite the absence of known anthropogenic Hg sources and major tributaries in this part of the lake catchment. Between the two sampling locations, the extensive delta of the Maggia river creates a zone where the width of the Lake is halved compared to the other part of the basin (Fig. 1). This morphological feature may isolate
the northernmost part of the Lake (site 29) from the rest of the basin and limit Hg transport from sources located further south (see below).

Mercury concentrations further increased moving south (Fig. 5). In the central part of the Lake, Hg concentrations were 0.202 ± 0.037 and 0.626 ± 0.229 mg kg⁻¹ in cores 54 and 55, respectively. Both values were significantly higher than Hg levels in the northern part of the Lake and the difference between sites 54 and 55 was also significant (Table S14). Mercury levels at site 55 may reflect the presence of historical, local pollution sources. A hat factory has been in operation in the town of Ghiffa (Fig. 1) from 1881 to 1981 and used Hg to cure pelts in the hat making process during part of this period. In 2008, during low-water level conditions, the presence of patches of metallic mercury close to the shoreline was brought to the attention of the Regional (Piedmont) Environmental Agency. A survey was promptly carried out, revealing Hg levels up to 26 mg kg⁻¹ (Fig. S1, Table S15). Although the highest concentrations were measured close to the coastline at depths < 50 m, concentrations between 1.5 and 3.6 mg kg⁻¹ occurred at depths up to 300 m (Fig. S1, Table S15, sampling points 3 to 6) and also south of the town of Ghiffa (sampling point 22 in Fig. S1).

Alternatively, we note that site 55 is located in the zone of maximum depth of the lake basin (370 m) and might be impacted by inputs of contaminated material carried out of the Borromean Gulf during major floods of river Toce. According to local knowledge, the flood plume of suspended matter originating from river Toce can be seen to stretch outside the Borromean Gulf up to the locality of Verbania-Intra and Ghiffa close to site 55 (Fig. 1). A major flood occurred in the River Toce in the year 2000 and its impact can be clearly seen in the lowest sections of cores collected at site 17 in the Borromean Gulf (Fig. 5). Although the general trend of water transport in Lake Maggiore consists in a southward movement (section 3.2), Ambrosetti et al. (2012) showed that soluble tracers can also move northwards at depth higher than 120–130 m. Deep waters from the Borromean Gulf (maximum depth = 120 m) could therefore move northwards after entering the central part of the lake where depth above 150 m are found at only 500 m from the lakeshore. The transport of Hg-contaminated material from the Borromean Gulf to site 55 (and possibly 54) cannot therefore be excluded and more detailed studies are required on the dynamics of Hg transport at the basin scale.

<figure><img src="image_8.png" /></figure>

Figure 5. Panel A. Total mercury concentrations in sediment sections of cores collected in 2008 at different sites along the north-south axis (N---S; main basin of Lake Maggiore) and in the Borromean Gulf (see Fig. 1 for sampling locations). Cores from different locations were divided in section with variable thickness (Table S6) to obtain temporally comparable snapshots of Hg contamination. Panel B. Same as panel A, but for cores collected in 2011. Note the different scales on the y-axes for the Hg concentrations below the axis break.
742

743

744

745

746

747

748

749

750

751

752

753

754

In any case, locally high background Hg levels at site 55 are unlikely. Mercury levels of 0.071 mg kg⁻¹ (Table 3) were measured in a deep section (94–95 cm) of a core collected about 2 km north of site 55. This core (LM_06/10/52/1) was sampled in 2006 at coordinates 45° 57' 28" N, 8°38'6" E and a depth of 375 m (Marchetto et al., 2007). Although core 52 was not dated to this depth, a sedimentation rate of 0.46 cm yr⁻¹ was established for the period 1963–2006 using diatomological analysis (see section 2.3). This sedimentation rate is comparable with values obtained from cores sampled at site 55 in 2008 and 2011 (Table S1). Based on the available sedimentation rates, the bottom section of core 52 could correspond to calendar ages of 1593 AD, 1799 AD and 1857 AD for sedimentation rates of 0.23, 0.46 and 0.64 cm yr⁻¹, respectively. All estimates predate the entry into operation of the hat factory (1881 AD) and do not support the hypothesis of Hg background levels higher than 0.1 mg kg⁻¹ in the central part of the lake. As a cautionary note, we acknowledge that assuming constant sedimentation rates neglects the possible influence of turbidite deposits associated to River Toce floods.

755

756

757

758

759

760

761

The dynamics of Hg contamination in cores from sites 27 (0.478 ± 0.059 mg kg⁻¹) and 28 (1.49 ± 0.318 mg kg⁻¹) has been extensively discussed in sections 3.2.2 and 3.2.4. Concentrations at site 28 were significantly higher than at all other sites (Table S14), including the highly impacted sites 17 and 51 in the Borromean Gulf. Mercury levels at site 27 were also significantly higher than at sites 29, 30b, 54 and 16, but not different from those at sites 55, 17 and 51 (Table S14). These observations confirm that the southernmost part of the Lake remained the most affected by legacy Hg contamination in 2008 and 2011 (Fig. 5).

762

763

764

765

766

767

768

769

770

771

772

773

Average Hg concentrations in cores 17, 16 and 51 collected from the Borromean Gulf were comparable to those measured in the central and southern parts of the lake. The bottom section of core 17 was characterised by high Hg concentrations on both sampling occasions and a similar pattern was observed in core 51 in 2011 (Fig. 5). These sediments contained turbiditic material transported from the river Toce during major floods that occurred in 2000 and 2020 (river flow during floods > 2600 m³ s⁻¹; average river flow = 70 m³ s⁻¹). Considering the average Hg levels of 0.18 ± 0.2 mg kg⁻¹ in the < 63 µm sediment fraction of the river Toce for the period 2001–2018 (Marziali et al., 2021a), the flood likely remobilized Hg-rich material from contaminated soils or depositional zones in the river Toce catchment. On the other hand, Hg levels in core 16 were significantly lower than in cores 51 and 17 (Table S14). These observations agree with Hg contamination showing a strong spatial heterogeneity in the Borromean Gulf (Guzzella et al., 2002).

774

775

776

777

778

779

780

Post-2011 monitoring data were also available for three sites in the Borromean Gulf and for site LM28 (Table S11). Considering the uppermost sediment layer at each sampling date, Hg concentrations followed a general decreasing trend at all sites with levels in 2023 often being less than a half than those measured in 2014 (Fig. S2). As of 2023, the highest Hg levels still occurred in the southern part of the lake (site LM28), in agreement with earlier observations and confirming the need for continued monitoring in this part of the lake.

781

### 3.4. Potential sources of Hg: MC-ICP-MS analyses

782

783

784

785

786

787

While our analyses discovered variations in δ²⁰²Hg values, Δ¹⁹⁹Hg and Δ²⁰¹Hg values were indistinguishable from zero (Table S16), suggesting no discernible mass independent isotope fractionation (MIF) in Lake Maggiore sediments. Hence, the following discussion concentrates on δ²⁰²Hg data. Despite the limited number of samples, MC-ICP-MS results seem to further support the hypothesis of ‘industrial Hg’ being transported outside the Borromean Gulf and accumulating in the southern part of Lake Maggiore during the 20th century.

788

789

790

791

In core ML06_ex13A, the δ²⁰²Hg was much less negative in section 32–33 cm (absolute maximum of Hg contamination in the Borromean Gulf) than in section 106–107 (representative of natural Hg background levels) (Table 2). Section 32–33 certainly reflects the isotopic signature of the mercury used as a catalyst in the chlor-alkali plant of Pieve Vergonte. Otherwise
792

stated, the industrial mercury had a clearly distinct isotopic signature from the mercury naturally occurring in the rocks of the River Toce catchment that formed the uncontaminated sediment layer of the Borromean Gulf.

793

794

795

796

<caption>Table 2. Overview of the section analysed by MC-ICP-MS. Replicate measurements are provided where available.</caption>

797

<table><thead><tr><th>Sample ID</th><th>Depth (cm)</th><th>Hg (mg kg<sup>-1</sup>)</th><th>Sedimentation Period</th><th>δ<sup>202</sup>Hg</th></tr></thead><tbody><tr><td>LM06_ex13A</td><td>0–1</td><td>0.4</td><td>Recent sediment (2006 AD)</td><td>-0.60</td></tr><tr><td>LM06_ex13A</td><td>32–33</td><td>25.9</td><td>Hg contamination (1944-1950 AD)</td><td>-0.18</td></tr><tr><td>LM06_ex13A</td><td>44–45</td><td>1.85</td><td>Gold mining (1800-1850 AD)</td><td>-0.85</td></tr><tr><td>LM06_ex13A</td><td>106–107</td><td>0.082</td><td>Background (1300 AD)</td><td>-0.84 / -0.86</td></tr><tr><td>LM06_27</td><td>0–1</td><td>0.782</td><td>Recent sediment (2006 AD)</td><td>-0.99</td></tr><tr><td>LM06_27</td><td>11–12</td><td>17.2</td><td>Hg contamination (1944-1950 AD)</td><td>-0.29 / -0.3</td></tr><tr><td>LM08_28</td><td>0–0.3</td><td>1.54</td><td>Recent sediment (2008 AD)</td><td>-0.85</td></tr><tr><td>LM11_28</td><td>13.5–14.5</td><td>13.8</td><td>Hg contamination (1944-1950 AD)</td><td>-0.69 / -0.55</td></tr><tr><td>LM06/52</td><td>94–95</td><td>0.071</td><td>Background (no age available)</td><td>-0.72 / -0.64</td></tr><tr><td>LM08_55</td><td>0–0.5</td><td>0.62</td><td>Recent sediment (2008 AD)</td><td>-0.70</td></tr></tbody></table>

798

799

800

801

802

803

804

805

806

807

808

809

810

811

812

813

814

815

816

817

818

819

820

821

822

823

824

In the same core, the δ<sup>202</sup>Hg of section 44–45 cm (containing Hg inputs from industrial gold mining in the River Toce basin, Fig. 2) was indistinguishable from that of the section 106–107 cm. In our opinion, it is likely that the mercury used to amalgam gold had an isotopic signature similar to the natural Hg in the rocks of the River Toce catchment although we are unable to verify this hypothesis. Finally, the δ<sup>202</sup>Hg signature in the topmost section of core LM06_ex13a still held a clear memory of industrial Hg sources in 2006, albeit suggesting a progressive dilution of heavier ‘industrial Hg’ with lighter natural Hg during the recovery phase (Table 2).

Outside the Borromean Gulf, cores LM06_27 and LM11_28 had less negative δ<sup>202</sup>Hg signatures in the sections corresponding to peak Hg contamination (section 11–12 cm and 13.5–14.5 cm, respectively) than in recent sediments. These data further support the hypothesis of Hg from the chlor-alkali plant being transported southwards at least during the period of maximum industrial Hg inputs. On the other hand, δ<sup>202</sup>Hg values in surface sediments were comparable with (LM11_28), or lower than (LM06_27), those of the natural background Hg concentrations of core LM06_ex13a. The high Hg levels in recent sediments from the southern part of the lake suggest a persistent legacy Hg contamination of industrial origin, while the heavier isotopic Hg signatures (lower δ<sup>202</sup>Hg values) would point to a natural origin of the residual Hg.

The section from cores LM08_55 and LM06_52 had similar δ<sup>202</sup>Hg signatures despite a 10-fold difference in total Hg levels. The more recent sediments of core LM08_55 were likely a mixture of material with Hg background concentrations (similar to those observed for core LM06_52) and of mercury used at the hat factory in Ghiffa (Fig. 1). Because the samples in our dataset did not allow to infer endmember signatures for different Hg sources, the actual origin of anthropogenic Hg in the central part of Lake Maggiore requires further research, especially in relation to the water circulation patterns discussed in section 3.2.
## 4. Conclusions

The results of the present study collectively support the hypothesis that Hg inputs discharged from a single industrial source affected the Borromean Gulf and the southern part of Lake Maggiore during most of the 20th century. In particular, stable isotope analyses using MC-ICP-MS confirms the existence of an anthropogenic Hg pool with a specific isotopic signature being discharged from a chlor-alkali plant. As of AD 2006, isotopic signatures more similar to those of the natural Hg pool indicate a progressive dilution of anthropogenic Hg, confirming the efficiency of remedial actions put in place to avoid further Hg export from the historical industrial point source. The recovery trend is confirmed by further monitoring studies between 2014 and 2023. We note however that anthropogenic mercury likely moved farther south than the zone of concern (SIN) delimited in 2000 by the Italian Ministry of the environment.

The detailed core chronologies allowed to confidently estimate the background Hg concentrations at around or below 0.1 mg kg⁻¹ in the whole lake. Concentrations above this level or, more importantly, exceeding levels of potential ecotoxicological concern ought not to be ascribed to the putative presence of high background concentrations due to geological peculiarities in Lake Maggiore catchment, specifically in the Borromean Gulf.

Finally, analysis of Hg contamination at 1-cm resolution (or less) in sediment cores allowed to derive simple models to estimate the time for recovery from peak Hg contamination to background levels. According to these models, it will take at least until the 2040s to bring Hg concentrations back to natural levels at sites originally not directly impacted by Hg industrial inputs. The recovery period could be much longer in the Borromean Gulf. On the other hand, recent monitoring data suggests that models may slightly underestimate the speed of recovery and actual Hg concentrations are constantly responding to other factors such as the phase-out of Hg-catalytic cells or extreme meteo-climatic events. The latter include the extreme floods of the Toce River in 2000 and 2020 and the severe drought conditions registered in 2021–2023 in the whole lake basin.

In terms of human and environmental health, Hg risk is driven by the strong bioaccumulation and biomagnification potential of methylmercury; issues that were not examined in the present study. However, the present paper shows that a thorough basin-scale paleolimnological approach can provide invaluable information on: i) the recovery time required for Hg concentration to fall back to background levels, ii) the possible internal processes controlling Hg circulation at the basin scale, and iii) the possible contribution of different Hg sources to the observed total Hg concentration in different parts of the basin.

Despite the presence of the legally binding Minamata Convention on Mercury (entered into force in 2017), legacy mercury contamination remains a global problem. Studies allowing to estimate the time for back-to-background Hg concentrations should be further encouraged to integrate long-term remediation operation with natural recovery processes in management decisions. Similarly, knowing the contribution of different sources allows to prioritize possible actions to effectively remediate Hg contamination in anthropogenically-impacted ecosystems.

## Supporting information

Data and supplementary information for the present article (16 tables and 2 figures) are available on-line at http://dx.doi.org/10.24396/ORDAR-193 (Vignati et al., 2025).

**Acknowledgements.** We thank Jean-Luc Loizeau (University of Geneva, Department F.-A. Forel for Environmental and Aquatic Sciences) for his assistance with Stokes' law calculations, Piero Guilizzoni (CNR-IRSA, Verbania, Italy) for his guidance in planning research and support in field laboratory work, and Prof. Chris Metcalfe (Trent University) for his help in facilitating access to MC-ICP-MS equipment.
875 Author contributions: CRediT

876 Davide A.L. Vignati: Conceptualization, Funding acquisition, Formal analysis, Visualization,

877 Data curation, Writing – original draft

878

879 Aldo Marchetto: Conceptualization, Investigation, Resources, Project administration, Writing

880 – original draft

881 Andrea Lami: Investigation, Resources, Visualization

882 Benedetta Ponti: Investigation

883 Camusso: Conceptualization, Funding acquisition, Supervision

884 Lucia Valsecchi: Investigation

885 Angela Boggero: Resources, Writing – review & editing

886 Simona Musazzi: Investigation, Resources, Visualization

887 Holger Hintelmann: Investigation, Resources, Writing – original draft

888 Roberta Bettinetti: Supervision, Writing – review & editing

889 Laura Marziali: Conceptualization, Investigation, Funding acquisition, Project administration,

890 Writing – review and editing, Supervision

891

892 Funding sources

893 This research was partially funded by the International Commission for the Protection of

894 Italian-Swiss Waters (CIPAIS), Research Programs 2013–2015, 2016–2018 and 2019–2021

895 (https://www.cipais.org/web/).

896

897

898 References

899 Agostinetti, P. P., et al., Gold in the ALPS: A View from the South. In: G. Morteani, J. P. Northover, Eds.),

900 Prehistoric Gold in Europe: Mines, Metallurgy and Manufacture. Springer Netherlands,

901 Dordrecht, 1995, pp. 199-218.

902 Ambrosetti, W., et al., Valutazione del tempo di residenza delle acque lacustri [Evaluation of residence

903 times of lake waters]. In: R. Bertoni, (Ed.), Ricerche sull'evoluzione del Lago Maggiore: Aspetti

904 limnologici [Investigations on the evolution of Lake Maggiore: limnological aspects]. CIPAIS,

905 Commissione Internazionale per la protezione delle acque italo-svizzere [International

906 commission for the protection of Italian-Swiss waters], Verbania Pallanza, Italy, 2009.

907 Ambrosetti, W., et al., 2012. Hydraulic paths and estimation of real residence time of the water in Lago

908 Maggiore (N. Italy): application of massless markers transported in 3D motion fields. Journal

909 of Limnology. 71, 23-33.

910 Bank, M. S., 2020. The mercury science-policy interface: History, evolution and progress of the

911 Minamata Convention. Science of The Total Environment. 722, 137832.

912 Barbanti, L., Ambrosetti, W., 1989. Geological and geographical features of the Lago Maggiore drainage

913 basin. Memorie dell'Istituto Italiano di Idrobiologia Dott. Marco De Marchi. 46, 1-20.

914 Baudo, R., et al., 1992. I sedimenti del Lago di Varese. 2. Distribuzione di elementi maggiori, metalli in

915 traccia ed elementi rari [The sediments of Lake Varese. 2. Distribution of major elements, trace

916 elements and rare elements]. Acqua & Aria. 5, 435-442.

917 Baudo, R., et al., 1981. Mercury levels in sediment from Southern Lake Maggiore. Mem. Ist. Ital.

918 Idrobiol. 38, 209-220.

919 Bertola, M. R., Bielli, E., Indagini sul comparto ittico. In: A. Calderoni, (Ed.), Monitoraggio della presenza

920 del DDT e di altri contaminanti nell'ecosistema Lago Maggiore [Monitoring of DDT and other

921 contaminants in the ecosystem of Lake Maggiore]. CIPAIS, Commissione Internazionale per la

922 protezione delle acque italo-svizzere [International commission for the protection of Italian-

923 Swiss waters], Verbania Pallanza, Italy, 2002, pp. 74-83.

924 Blum, J. D., Bergquist, B. A., 2007. Reporting of variations in the natural isotopic composition of

925 mercury. Analytical and Bioanalytical Chemistry. 388, 353-359.
926 Bolaños-Alvarez, Y., et al., 2024. Regional assessment of the historical trends of mercury in sediment cores from Wider Caribbean coastal environments. *Science of The Total Environment*. 920, 170609.

929 Calderoni, A., et al., Ricerche sulla distribuzione e gli effetti del DDT nell'ecosistema Lago Maggiore [Investigations on the distribution and effects of DDT in Lake Maggiore ecosystem]. *CIPAIS*, Commissione Internazionale per la protezione delle acque italo-svizzere [International commission for the protection of Italian-Swiss waters], *Verbania Pallanza*, Italy, 1999.

933 Camusso, M., Borsani, G., Indagine conoscitiva sui livelli di mercurio nei comparti ambientali del Lago Maggiore [Investigation on mercury levels in the environmental compartments of Lake Maggiore]. *Internal Report of CNR-IRSA*. 71pp., 2004.

936 Ciampittiello, M., 1999. I livelli del Lago Maggiore: una grande risorsa da gestire, un problema da affrontare. *Lions Club*, CNR, *Alberti Libraio Editore*, *Verbania-Intra*, Italy.

938 Covelli, S., et al., 2017. Historical sedimentary trends of mercury and other trace elements from two saltmarshes of the Marano and Grado lagoon (northern Adriatic Sea). *Journal of Soils and Sediments*. 17, 1972-1985.

941 Damiani, V., Thomas, R. L., 1974. Mercury in the sediments of Pallanza Basin. *Nature*. 251, 696-697.

942 Del Soldato, M., 1986. Giacimenti auriferi e cultura materiale in Valle Anzasca. *Boll. Ass. Min. Subalp.* 23, 499-503.

944 Del Soldato, M., 1989. La distillazione dell'amalgama nella metallugria dell'oro ossolana dei secoli XVII-XIX. *Boll. Ass. Min. Subalp.* 26, 73-76.

946 Fantozzi, L., et al., 2021. An Integrated Investigation of Atmospheric Gaseous Elemental Mercury Transport and Dispersion Around a Chlor-Alkali Plant in the Ossola Valley (Italian Central Alps). *Toxics*. 9, 172.

949 Fenocchi, A., et al., 2017. Relevance of inflows on the thermodynamic structure and on the modeling of a deep subalpine lake (Lake Maggiore, Northern Italy/Southern Switzerland). *Limnologica*. 63, 42-56.

952 Foucher, D., et al., 2009. Tracing Mercury Contamination from the Idrija Mining Region (Slovenia) to the Gulf of Trieste Using Hg Isotope Ratio Measurements. *Environmental Science & Technology*. 43, 33-39.

955 Giuliani, A., et al., 2015. Preliminary Studies Aimed at the Re-Opening of Pestarena Gold Mine (North-Western Alps, Italy). *American Journal of Environmental Sciences*. 11.

957 Grassi, S., La condizione dei lavoratori dello stabilimento chimico di Pieve Vergonte nella storia (1915-2012). Facoltà di Scienze dell'Informazione e dell'Economia. Università degli Studi di Modena e Reggio Emilia, Modena and Reggio Emilia, 2012, pp. 269.

960 Grassi, S., Rumianca - Pieve Vergonte (VB). https://www.industriaeambiente.it/schede/pieve_vergonte/. Last accessed on 02/09/2025., Vol. 2025, 2013.

963 Guilizzoni, P., Lami, A., 1986. Paleolimnology of the eutrophic Lake Varese (Northern Italy). 2. Concentration of ten heavy metals in a sediment core. *Memorie dell'Istituto Italiano di Idrobiologia Dott. Marco De Marchi*. 44, 15-26.

966 Guilizzoni, P., et al., 2012. Ecological effects of multiple stressors on a deep lake (Lago Maggiore, Italy) integrating neo and palaeolimnological approaches. *J. Limnol*. 71, 1-22.

968 Guzzella, L., et al., Contaminanti nei sedimenti del Lago Maggiore [Contaminants in sediments of Lake Maggiore]. In: A. Calderoni, (Ed.), *Monitoraggio della presenza del DDT e di altri contaminanti nell'ecosistema Lago Maggiore* [Monitoring of DDT and other contaminants in the ecosystem of Lake Maggiore]. *CIPAIS, Commissione Internazionale per la protezione delle acque italo-svizzere* [International commission for the protection of Italian-Swiss waters], *Verbania Pallanza*, Italy, 2002, pp. 19-46.

974 Guzzella, L. M., et al., 2018. Spatial and temporal trends of target organic and inorganic micropollutants in Lake Maggiore and Lake Lugano (Italian-Swiss water bodies): contamination in sediments and biota. *Hydrobiologia*. 824, 271-290.
977 Kämpf, L., et al., 2012. Detrital layers marking flood events in recent sediments of Lago Maggiore (N. Italy) and their comparison with instrumental data. Freshwater Biology. 57, 2076-2090.

978 Lacerda, L. D., 1997. Global mercury emissions from gold and silver mining. Water, Air, and Soil Pollution. 97, 209-221.

981 Lin, D., et al., 2017. Toolset for assessment of natural recovery from legacy contaminated sediment: Case study of Pallanza Bay, Lake Maggiore, Italy. Water Research. 121, 109-119.

983 Locht, B., Muntau, H., 1981. La composition chimique des sédiments profonds du Lac Majeur. Mem. Ist. Ital. Idrobiol. 38, 101-186.

985 MacDonald, D. D., et al., 2000. Development and evaluation of consensus-based sediment quality guidelines for freshwater ecosystems. Arch. Environ. Contam. Toxicol. 39, 20-31.

987 Madau, G., et al., 1986. Ipotesi di lavoro per la verifica delle potenzialità giacimentologiche nelle strutture filoniane di Pestarena [A working hypothesis to evaluate the ore deposit potential of the dyke structures at Pestarena]. Boll. Ass. Min. Subalp. 23, 505-534.

990 Mao, L., et al., 2023. Tracking the multiple Hg sources in sediments in a typical river-lake basin by isotope compositions and mixing models. Journal of Hazardous Materials. 459, 132166.

992 Marchetto, A., et al., Campionamento, preparazione dei campioni per le analisi di DDT e correlazione tra carote mediante studio diatomologico [Collection and preparation of samples for DDT analysis and correlation among cores using a diatomological approach]. In: P. Guilizzoni, A. Calderoni, Eds.), Monitoraggio della presenza del DDT e altri contaminanti nell'ecosistema Lago Maggiore [Monitoring the presence of DDT and other contaminants in Lake Maggiore ecosystem]. CIPAIS, Commissione Internazionale per la protezione delle acque italo-svizzere [International commission for the protection of Italian-Swiss waters], Verbania Pallanza, Italy, 2007.

1000 Marchetto, A., et al., 2004. Lake Maggiore (N. Italy) trophic history: fossil diatom, plant pigments, and chironomids, and comparison with long-term limnological data. Quaternary International. 113, 97-110.

1003 Marziali, L., et al., Concentrazioni di mercurio e altri elementi in traccia [Concentrations of mercury and other trace elements]. In: L. Marziali, (Ed.), Indagini sulle sostanze pericolose nell'ecosistema del Lago Maggiore [Investigations on dangerous substances in Lake Maggiore ecosystem]. CIPAIS, Commissione Internazionale per la protezione delle acque italo-svizzere [International commission for the protection of Italian-Swiss waters], Brugherio, 2024.

1008 Marziali, L., et al., 2021a. Twenty-year sediment contamination trends in some tributaries of Lake Maggiore (Northern Italy): relation with anthropogenic factors. Environmental Science and Pollution Research. 28, 38193-38208.

1011 Marziali, L., et al., 2022. Zooplankton as Mercury Repository in Lake Maggiore (Northern Italy): Biomass Composition and Stable Isotope Analysis. Water. 14, 680.

1013 Marziali, L., et al., 2021b. Mercury Bioaccumulation in Benthic Invertebrates: From Riverine Sediments to Higher Trophic Levels. Toxics. 9, 197.

1015 Marziali, L., et al., Misura delle concentrazioni di elementi in traccia nei sedimenti del lago. In: A. Marchetto, (Ed.), Indagini su DDT e sostanze pericolose nell'ecosistema del Lago Maggiore [Investigations on DDT and dangerous substances in Lake Maggiore ecosystem]. CIPAIS, Commissione Internazionale per la protezione delle acque italo-svizzere [International commission for the protection of Italian-Swiss waters], Verbania Pallanza, Italy, 2015.

1020 Marziali, L., et al., Sediment contamination in Lake Maggiore and its main tributaries: analysis of DDT, PBDEs, PCBs, PAHs, HBCD, DBDPE, synthetic fragrances, PFASs and trace elements (Hg, As, Cd, Cu, Ni, Pb). https://doi.org/10.5281/zenodo.16631896. 2025.

1023 Marziali, L., Valsecchi, L., Misura delle concentrazioni di mercurio e di altri elementi in traccia [Concentrations of mercury and other trace elements]. In: A. Marchetto, (Ed.), Indagini su DDT e sostanze pericolose nell'ecosistema del Lago Maggiore [Investigations on DDT and dangerous substances in Lake Maggiore ecosystem]. CIPAIS, Commissione Internazionale per la protezione delle acque italo-svizzere [International commission for the protection of Italian-Swiss waters], Verbania Pallanza, Italy, 2022.
1029 Marziali, L., et al., Misura delle concentrazioni di mercurio e di altri elementi in traccia nei sedimenti del lago [Concentrations of mercury and other trace elements in lake sediments]. In: A. Marchetto, (Ed.), Indagini su DDT e sostanze pericolose nell'ecosistema del Lago Maggiore [Investigations on DDT and dangerous substances in Lake Maggiore ecosystem]. CIPAIS, Commissione Internazionale per la protezione delle acque italo-svizzere [International commission for the protection of Italian-Swiss waters], Verbania Pallanza, Italy, 2018.

1035 Menaballi, A., Studio della contaminazione de mercurio in componenti abiotiche e biotiche del Lago Maggiore [Study of mercury contamination in abiotic and biotic components of Lake Maggiore]. Scienze matematiche fisiche e naturali; Scienze biologiche. Università degli Studi di Milano, Milan, Italy, 1999.

1039 Ministerodell'Ambiente, 2000. Perimetrazione del sito di interesse nazionale di Pieve Vergonte. Decreto 10 gennaio 2000. GU Serie Generale n.46 del 25-02-2000.

1041 Muntau, H., 1981. Heavy metal distribution in the aquatic ecosystem "Southern Lake Maggiore": II. Evaluation and trend analysis. Mem. Ist. Ital. Idrobiol. 38, 505-529.

1043 Putyrskaya, V., Klemt, E., 2007. Modeling 137Cs migration processes in lake sediments. Journal of Environmental Radioactivity. 96, 54-62.

1045 Rogora, M., et al., 2021. Temporal changes in nutrients in a deep oligomictic lake: the role of external loads versus climate change. Journal of Limnology. 80.

1047 Rossi, D., et al., 1993. Chemical composition of Lake Maggiore sediments in 1986. Mem. Ist. Ital. Idrobiol. 51, 53-114.

1049 Scapozza, C., Patocchi, N., 2023. Lake Maggiore: geomorphological genesis, lake-level evolution, and present and future ecosystems importance. Journal of Limnology. 81.

1051 Schudel, G., et al., 2018. An investigation of mercury sources in the Puyango-Tumbes River: Using stable Hg isotopes to characterize transboundary Hg pollution. Chemosphere. 202, 777-787.

1053 Selin, N. E., 2009. Global Biogeochemical Cycling of Mercury: A Review. Annual Review of Environment and Resources. 34, 43-63.

1055 Sun, R., et al., 2013. Mercury stable isotope fractionation in six utility boilers of two large coal-fired power plants. Chemical Geology. 336, 103-111.

1057 Sunderland, E. M., Selin, N. E., 2013. Future trends in environmental mercury concentrations: implications for prevention strategies. Environmental Health. 12, 2.

1059 USEPA, 1998. SW-846 Test Method 7473: Mercury in Solids and Solutions by Thermal Decomposition, Amalgamation, and Atomic Absorption Spectrophotometry. Revision 0. US Environmental Protection Agency, Washington, DC.

1062 Varekamp, J., et al., 2003. Mercury contamination chronologies from Connecticut wetlands and Long Island Sound sediments. Environmental Geology. 43, 268-282.

1064 VariousAuthors, Lago Maggiore [Lake Maggiore]. Memorie dell'Istituto Italiano di Idrobiologia Dott. Marco De Marchi, 1989, pp. 324pp.

1066 Viel, M., Damiani, V., 1985. Sedimentological and geochemical study of recent deposits in Lago Maggiore (North Italy). Mem. Ist. Ital. Idrobiol. 43, 181-238.

1068 Vignati, D. A. L., et al., 2020. Testing the Use of Standardized Laboratory Tests to Infer Hg Bioaccumulation in Indigenous Benthic Organisms of Lake Maggiore (NW Italy). Applied Sciences. 10, 1970.

1071 Vignati, D. A. L., Guilizzoni, P., 2011. Metalli nel Lago Maggiore: livelli naturali e antropici. Acqua & Aria. 1, 22-27.

1073 Vignati, D. A. L., et al., Contaminanti nei sedimenti del Lago Maggiore. Mercurio totale [Contaminants in sediments of Lake Maggiore. Total mercury]. In: P. Guilizzoni, (Ed.), Indagini su DDT e sostanze pericolose nell'ecosistema del Lago Maggiore [Investigations on DDT and dangerous substances in Lake Maggiore ecosystem]. CIPAIS, Commissione Internazionale per la protezione delle acque italo-svizzere [International commission for the protection of Italian-Swiss waters], Verbania Pallanza, Italy, 2012, pp. 29-32.
1079 Vijayaraghavan, K., et al., 2014. Response of fish tissue mercury in a freshwater lake to local, regional, and global changes in mercury emissions. Environmental Toxicology and Chemistry. 33, 1238-1247.

1082 Wang, F., et al., 2019. How closely do mercury trends in fish and other aquatic wildlife track those in the atmosphere? – Implications for evaluating the effectiveness of the Minamata Convention. Science of The Total Environment. 674, 58-70.

1085 Wiederhold, J. G., et al., 2015. Mercury Isotope Signatures in Contaminated Sediments as a Tracer for Local Industrial Pollution Sources. Environmental Science & Technology. 49, 177-185.