Recent history of persistent organic pollutants (PAHs, PCBs, PBDEs) in sediments from a large tropical lake

Accepted Manuscript Title: Recent history of persistent organic pollutants (PAHs, PCBs, PBDEs) in sediments from a large tropical lake Authors: Jorge Feliciano Ontiveros-Cuadras, Ana Carolina Ruiz-Fern´andez, Joan Albert Sanchez-Cabeza, Jos´e Sericano, Libia Hascibe P´erez-Bernal, Federico P´aez-Osuna, Robert B. Dunbar, David A. Mucciarone PII: DOI: Reference:

S0304-3894(18)31029-X https://doi.org/10.1016/j.jhazmat.2018.11.010 HAZMAT 19932

To appear in:

Journal of Hazardous Materials

Received date: Revised date: Accepted date:

16 February 2018 21 October 2018 4 November 2018

Please cite this article as: Ontiveros-Cuadras JF, Ruiz-Fern´andez AC, SanchezCabeza JA, Sericano J, P´erez-Bernal LH, P´aez-Osuna F, Dunbar RB, Mucciarone DA, Recent history of persistent organic pollutants (PAHs, PCBs, PBDEs) in sediments from a large tropical lake, Journal of Hazardous Materials (2018), https://doi.org/10.1016/j.jhazmat.2018.11.010 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

Recent history of persistent organic pollutants (PAHs, PCBs, PBDEs) in sediments from a large tropical lake

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Ontiveros-Cuadras, Jorge Felicianoa, Ruiz-Fernández, Ana Carolinab*, Sanchez-Cabeza, Joan Alberta, Sericano, Joséc, Pérez-Bernal, Libia Hascibeb, Páez-Osuna, Federicob, Dunbar, Robert B.d, Mucciarone, David A.d

a

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Unidad Académica Procesos Oceánicos y Costeros, Instituto de Ciencias del Mar y Limnología,

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Universidad Nacional Autónoma de México, Ciudad Universitaria, 04510, Ciudad de México. E-mail:

Unidad Académica Mazatlán, Instituto de Ciencias del Mar y Limnología, Universidad Nacional

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b

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[email protected], [email protected]

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Autónoma de México, 82040, Mazatlán, México. E-mail: [email protected], lbernal@

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ola.icmyl.unam.mx, [email protected] c

Geochemical and Environmental Research Group, Texas A&M University, 833 Graham Road,

Earth System Science, Stanford University, Stanford, CA94305 USA. E-mail: [email protected],

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d

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College Station, TX 77845, USA. E-mail: [email protected]

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[email protected]

*

Corresponding author: [email protected]

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Highlights  PAH, PCB and PBDE sediment concentrations were studied in Chapala Lake, Mexico  POP sediment concentrations were indicative of moderate to intense contamination  POP sediment concentrations can be harmful for human through fish consumption

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 Despite the banning of POP use, sediment concentrations are still growing since 1990s  Long-range atmospheric transport has been the main source of POPs to the sediments

Abstract

Pb-dated sediment cores and surface sediments from Lake Chapala (LC), Mexico, were analyzed to

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assess the temporal trends in concentrations and fluxes of persistent organic pollutants (POPs: PAHs,

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PCBs and PBDEs). Total sediment concentrations of PAHs (95-1482 ng g-1), PCBs (9-27 ng g-1) and

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PBDEs (0.2-2.5 ng g-1) were indicative of moderate to intense contamination. The POP concentrations have progressively increased since the 1990s. The light molecular weight PAHs, and the prevalence of

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PCB congeners with low-chlorination levels (e.g., di- to tri-CB) and low-to medium-brominated (tri- to

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penta-BDE) PBDEs in most sections of the sediment profiles, suggested that these POPs have most

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likely reached these sediments by long-range atmospheric transport from distant sources; although the significant presence of heavier PAH, PCB and PBDE congeners in the topmost sediments, indicate that

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other nearby and local sources (soil erosion from the catchment, urban and industrial wastewaters discharges, as well as navigation) might have also contributed to the recent input of POPs to LC.

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Taking into account the relevance of LC as regional freshwater supply and commercial fishing ground, the potential risk posed by the organic contaminated sediments to the biota and human population should not be underestimated.

Keywords:

PAHs,

PCBs,

PBDEs,

Lake

sediments,

Long

range

atmospheric

transport 2

1. Introduction Lacustrine sediments are often seen as the final repository of materials that come from land, the atmosphere, and terrestrial waters. By extension, when sediment perturbation within a lake has been low, a sediment core with a reliable age-depth model (i.e.

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Pb radiochronology) becomes an

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environmental archive allowing the reconstruction of a trustworthy picture of present and past conditions in the lacustrine system [1, 2].

Persistent organic pollutants (POPs) are ubiquitous contaminants with a multiplicity of sources and transport mechanisms; they have been extensively used, especially since the end of World War II. Most POPs are volatile and might recycle among air, water, and soil at ordinary environmental

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temperatures, with warmer temperatures promoting evaporation and atmospheric dispersion through

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long distances, and cooler temperatures allowing atmospheric deposition onto soil and water [3].

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Lacustrine sediments might constitute a repository of POPs such as PAHs (polycyclic aromatic

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hydrocarbons), PCBs (polychlorinated biphenyls) and PBDEs (polybrominated diphenyl ethers) owing

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to their strong affinity for particulate matter and their slow degradation rates, especially in anoxic and

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light-protected environments [2, 4]. POPs photodegradation depends on the structure of the homologue degraded [5]; e.g. phenanthrene (Phe) and dibenzo(a,h)anthracene (DahAn), PAH homologues with

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structurally stable molecules, degrade slowly under direct light [6], whereas highly chlorinated PCB

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congeners are more readily dechlorinated than lower chlorinated congeners [7] . The toxicity and persistence of POPs, as well as their lipophilic capacity, potentiate bioaccumulation and

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biomagnification in wildlife [8, 9] and in humans. Some adverse effects of POPs and their metabolites in human populations include neurotoxicity, immunosuppression, endocrine system damage, reproductive dysfunction and carcinogenicity [10, 11]. In Mexico, historical data of POPs in environmental matrices are sparse. Some studies have recorded PAHs and PCBs in sediments from the Coatzacoalcos River on the Gulf of Mexico [12], the 3

Colorado River in Mexicali Valley [13], coastal lagoons on the Gulf of Mexico [14,15] or the Caribbean Sea [16]; and lacustrine systems on the Central Mexican Plateau [7, 17] and in urban areas around Mexico City [18, 19]. LC is the largest natural freshwater reservoir in Mexico (~1,100 km2 and 9,686 Mm3 storage

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capacity, Fig. 1). It hosts a high diversity of endemic, some endangered, species and migratory birds (Ramsar site since 2009). It is the main freshwater supply for Guadalajara City (~8 million inhabitants [20]) and the surrounding industrial parks and agriculture fields, which has caused dramatic changes in the LC water budget due to water extraction. Guadalajara City is characterized by an accelerated population growth since 1910s and an early industrialization process since the 1930s, which

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consolidated between 1950-1960, a period in which also the population practically doubled [21].

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Guadalajara City metropolitan area has experienced ~48% population growth between 1990 and 2010,

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and an increase in water demand rate (~5.6% per year) as a result of the expansion of its industrial

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sector (~2.9% annual average growth during the period; mainly textile, metal-mechanical and food

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production) [22, 23, 24, 25]. In the last decades, the LC water quality has been affected by runoff from

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agricultural fields and untreated industrial and domestic wastewater discharges [26]. Previous studies reported high Hg and other trace element concentrations in fish (e.g., whitefish Coregonus spp., blue

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tilapia Oreochromis aureus) and sediments [27], which were related to soil erosion in the watershed

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[28].

Our work aimed to reconstruct the long-term accumulation trends of POPs in the largest

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lacustrine system of Mexico, under the hypothesis that POP concentrations have increased during the past as a result of the regional socioeconomic development, and should have decreased owing to POP use banning, and implementation of environmental regulations. The

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Pb-dated sediment cores

collected in LC were used to (1) define baseline and recent PAH, PCB, and PBDE sediment concentrations, (2) evaluate the potential ecological risk that recent POP concentrations might pose to 4

the benthic biota; and (3) reconstruct the historical trends of POP concentrations and fluxes within the length of the cores. In addition, a chemometric approach was used to elucidate possible sources and transport mechanisms that promote the presence of these POPs in LC sediments.

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2. Study area LC is in the western part of the Trans-Mexican Volcanic Belt, at 1,524 m a.s.l. It belongs to LermaSantiago-Pacifico watershed (drainage area ~140,000 km2), the most densely developed river basin in Mexico. It is a shallow system (mean depth ~7 m), surrounded by a catchment area of 54,000 km2 including croplands and forest exploitation, as well as industrial and urban developments [29]. Its main

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water source is the Lerma River (annual flow of 68.2 m3 s-1) which has serious water quality issues due

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to the extensive industrial development in its 700 km channel and, in also receives 2.4 m3 s-1 of

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wastewaters into the main channel at the northeastern side of the lake [30, 31]. The natural outflow is

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the Grande de Santiago River, draining directly into the Pacific Ocean [30]. Major anthropogenic

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modifications that affected the LC hydrological balance include the building of an earth embankment

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dam in 1910 which reduced LC surface area in ~30%, and the construction of dams along the Lerma River from 1950 to 1960 [32]. LC is a neotectonic lake with active hydrothermal activity. Chapala

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basin belongs to the western sector of the Pliocene-Quaternary Mexican Volcanic Belt [33] related to

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the motion of Cocos Plate relative to the North America Plate [34]. The regional climate is subtropical semi-arid with summer rainfalls [26]. Chapala basin is

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bounded by ranges (elevation difference between the top of the mountains and the lacustrine plain is ~500 m). Controlled by topography, the prevailing winds blow in west-east direction, generating currents and maelstroms in the lake [35]. The commercial fishing in LC accounted for 8,000 ton in 2010, including several native species of Chirostoma spp. (charal and whitefish) and Ictalurus spp. (catfish), as well as introduced species Cyprinus spp. (carp) and Oreochromis aureus (tilapia) [36]. 5

3. Material and methods 3.1. Sampling and laboratory analysis Four sediment cores were collected in LC using an UWITEC™ gravity corer (Plexiglas liner; 8.5 cm inner diameter): C1, C2 and C3 in July 2010 and C4 in February 2012 (Table 1). Neither laminations

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nor signs of bioturbation were observed along the sediment sequences. Cores were extruded and subsampled at 1 cm intervals; sediment samples were freeze-dried and ground to powder with an agate mortar and pestle, excepting the aliquots for grain size analysis. The POP concentrations reported correspond to the surface of cores C1 and C2 (the rest of the samples were not available), and along the length of cores C3 and C4.

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Grain size distribution was analyzed by laser diffraction (Malvern-Mastersizer 2000) using

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sediment aliquots previously treated with 30% H2O2. Magnetic susceptibility measurements were

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performed in a Bartington MS2 meter, using the Bartington-G039 calibration standard (accuracy 98%,

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reproducibility 4%). Organic matter concentrations were estimated by a volumetric procedure [38] with

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a reproducibility of ~5%. Bulk elemental and isotopic forms of carbon and nitrogen were measured in a

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Carlo Erba NA1500 Series 2 EA coupled to a Finnigan Delta Plus isotope ratio mass spectrometer via a Conflo II open-split interface at Stanford University. δ13C is reported relative to Vienna Pee Dee

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Belemnite (V-PDB) and δ15N to atmospheric N2. Sample reproducibility base in duplicate analyses was 0.05% for δ13C, 0.14% for wt % total organic C, 0.20% for δ15N, and 0.01% for wt % total N. Pb chronologies of cores C1, C2 and C3 have already been published elsewhere [28, 39,

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Pb activity was determined by alpha spectrometry (Ortec–Ametek 920E) assuming secular

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40].

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equilibrium with its radioactive descendant 210Po (115% accuracy and 3.5% reproducibility determined through replicate analyses of the certified material IAEA-300). Excess 210Pb (210Pbex) was calculated as the difference between total (210Pbtot) and supported

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Pb (210Pbsup) activities with

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Pbsup values

confirmed by 226Ra mean activities (Fig. 2a-d). 226Ra was determined in an ultralow background liquid 6

scintillation system (Quantulus 122TM) using alpha/beta discrimination [41] (114% accuracy, checked against the certified material IAEA-384).

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Pb chronologies, sediment accumulation rates (SAR, cm

yr-1) and mass accumulation rates (MAR, g cm-2 yr-1) for cores C1, C2 and C3 were calculated through the Constant Flux (CF) model [1, 42]; and for C4, through the Constant Flux Constant Sedimentation

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model (CFCS; [43]) (detailed explanation in section 4.1.). The almost identical trends in magnetic susceptibility (Fig. 1S) were used for stratigraphic correlation among cores C1, C2 and C3. And at core C2 the

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Pb-derived dates were corroborated using the weapon fallout tracer

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Pu (Fig. 2d) [39],

which was purified using ion interchange resins, isolated by electroplating and determined by alpha spectrometry [44].

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PAHs, PCBs and PBDs concentrations were analyzed by gas chromatography/mass spectrometry

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(GC/MS, Hewlett-Packard 5890/5970) at Texas A&M University, as described in [7] (details available

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as supplementary information). Briefly, samples were spiked with surrogate standards for each

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contaminant group, then extracted with dichloromethane by accelerated solvent extraction (Dionex

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ASE-200) and purified in a chromatography column (2:1 of silica-alumina). Deuterated aromatic

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compounds were used as recovery standards [7, 45]. Analysis of the standard reference material (NIST-

±25%).

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1944) were within the certified ranges and replicate sample analysis varied within acceptable limits (<

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The total PAH (ΣPAHs) concentrations included: (1) the low molecular weight PAHs (LMW PAHs, 3 or fewer benzene rings) naphthalene (Na), acenaphthylene (Acy), acenaphthene (Ace),

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fluorene (Fl), phenanthrene (Phe), and anthracene (An); and (2) the high molecular weight PAHs (HMW PAHs, 4 or more benzene rings) fluoranthene (Flt), pyrene (Py), benzo(a)anthracene (BaAn), chrysene (Ch), benzo(g,h,i)perylene (BghiPe) and the penta-aromatic congeners (Ʃpenta-PAHs) benzo(b)fluoranthene (BbFlt), benzo(k)fluoranthene (BkFlt), benzo(a)pyrene (BaPy), indeno(1,2,3cd)pyrene (Ipy) and dibenzo(a,h)anthracene (DahAn). BbFlt and BkFlt are reported as a single species 7

(B(b+k)Flt). Perylene (Pe) and benzo(e)pyrene (BePy) were also determined to assess PAH origins. PCB congeners analyzed were mono- to deca-chlorobiphenyls, and PBDE congeners were BDE 1, 2, 3, 7, 8, 10, 11, 12, 13, 15, 17, 25, 28, 30, 32, 33, 35, 37, 47, 49, 66, 71, 75, 77, 85, 99, 100, 116, 118, 119,

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126, 138, 153, 154, 155, 166, 181, 183, and 190.

3.2. Statistical analysis

A principal component analysis (PCA) was done with the XLSTAT-Pro 7.5 software, with VARIMAX rotation and factor adjustment of 2. The analysis included the PCB homologue concentrations from sediment cores C3 and C4 and from the superficial samples of cores C1 and C2, along with the PCB

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composition of the most commonly used Aroclor formulations in North America.

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4. Results and Discussion

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Pb activities (Fig. 2) in cores C1 (26.3 - 92.5 Bq kg-1), C2 (26.7 - 94.5 Bq kg-1) and C3

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The total

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4.1. Sediment chronology

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(27.0 - 84.3 Bq kg-1) were comparable with an almost constant value towards the core bottoms (31±3 34±3 Bq kg-1). Thus, the

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Pbsup values were determined as the average of

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Pb activity at those

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bottom sections [46]. The 210Pb depth profiles showed a departure from a monotonic exponential decay,

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which was interpreted as a result of sediment flux variations, for which the CF dating model is the most appropriate. In core C4, the

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Pb values were lower (37.9 - 63.2 Bq kg-1) although

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Pbsup activities

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(32.2±2.7 Bq kg-1) were comparable to the other three cores. The total 210Pb depth profile never reached the supported value, which indicated that the

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Pb depth profile and the

incomplete, precluding the use of the CF model for the

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210

Pb inventory were

Pb chronology computation. However, a

significant correlation (p < 0.05, r = 0.85) between the logarithm of 210Pbex and sediment mass depth (g cm-2) confirmed that the Constant Flux-Constant Sedimentation (CFCS) model could be used [1]. The 8

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Pb-dated portion of each core extends back to 95±2 yr for C1, 84±17 yr for C2, 92±13 yr for C3 and

44±4 yr for C4. The SAR and MAR were somewhat slower on the western side of LC (C1: 0.07 - 1.56 cm yr-1, 0.01 - 0.19 g cm-2 yr-1; C2: 0.24 - 2.26 cm yr-1, 0.05 - 0.15 g cm-2 yr-1) and increased in the central part of the lake, where two permanent gyres are developed (C3: 0.08 - 2.82 cm yr-1, 0.01 - 0.28

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g cm-2 yr-1) and at the mouth of the Lerma River (C4: 1.08±0.09 cm yr-1, 0.34±0.03 g cm-2 yr-1) as a consequence of hydrodynamic circulation [35].

4.2. Geochemical characterization

The sediments were composed mainly of silts (77 - 87%) and clays (8 - 19%). Magnetic susceptibility

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(MS) depth profiles were similar among cores C1, C2 and C3 (Fig. 1S), displaying an identical peak

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between the depths of 69 to 77 cm, followed by a return to basal levels in the

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Pb-dated portions.

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Grain size distribution and MS values varied little along the cores within the past century (Fig. 3)

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reflecting that, despite the land use changes or damming occurred during the past century, the source of

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detrital particles to LC have not changed significantly. Organic matter content in cores C3 and C4 was

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small (1 to 3%) but increased toward the sediment surface; the high C/N ratios (≥ 15) and the profiles of δ13C (< -26‰) and δ15N (4 - 9‰) also indicated a slight upward increase in the contribution of

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terrestrial organic matter to the lake [47].

4.3. Persistent organic pollutants

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4.3.1. PAHs

Total PAH (∑PAH) concentrations (Table 2) ranged from 95 to 1482 ng g-1 in core C3 and 360 to 661 ng g-1 in core C4, with maximum concentrations at the surface sections (Fig. 4). Ipy and DahAn were detected only in core C4, whereas BghiPe was not detected in core C3. The ∑PAH concentrations (Table 3) were higher than those reported for other rural and slightly contaminated urban sediments in 9

Mexico [7, 14, 17, 18, 19], but similar to values reported in slightly contaminated lakes in the Arctic [48], the urban lake Taihu in China [49] and the Mexican lakes Las Matas [15] and El Tule [7]. The ∑PAH concentrations in LC accounted for contaminated sediments (∑PAH > 100 ng g-1, [50]), although they were below the threshold effect concentration benchmark (TEC, 1610 ng g-1; [51]).

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Low-molecular-weight (LMW) compounds dominated the PAHs ensemble in surface samples C1 and C2 and in sediment cores C3 and C4, excepting in its topmost sections (Table 2; Fig. 5). LMW congeners such as Na, Phe and Fl are predominantly more volatile and are easily subject to long-range transport by winds [50]. The LMW PAHs surpassed the Threshold Effects Levels (TEL) value in the surface layer of core C3 [51], implying possible adverse effects on biota, and to humans through fish

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consumption. PAHs in sediments have been linked with liver neoplasms and other abnormalities in

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swimming activity, and DNA damage [54].

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bottom-dwelling fish [52, 53], and with delayed hatching, induction of deformities, disruption of larvae

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High-molecular-weight PAHs at the core top of C3 included Flt, Py and Ch, while in the surface

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of core C4 the most abundant species were B(b + k)Flt, Pe and BePy. In general, ∑PAH fluxes were

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lower in core C3 (6 - 235 ng cm-2 yr-1) than in C4 (122 - 225 ng cm-2 yr-1) which had a higher content of clays; and the maximum values were observed at the top of each sediment core (Fig. 4). In surface

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sediments (C1 and C2) and along the core C3, the Phe/Ant ratios were < 10, indicating a dominant

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PAH pyrogenic source; which could be related to the combustion of heavy fuel (An/(An+Phe) ratio, > 0.10) and fossil fuel, such as vehicle and crude oil (Flt/(Flt+Py), 0.4 - 0.5). However, these samples

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also showed low ratios of BaAn/Ch (< 0.4) and BaAn/(BaAn+Ch) (< 0.35) which indicated that PAHs were derived from a petrogenic source [55, 56, 57]. Similarly, in core C4, the ratios Phe/Ant, An/(An+Phe) and Flt/(Flt+Py) indicated that the PAHs in the sediments resulted from heavy and fossil fuel combustion; whereas the low values of BaAn/Ch and BaAn/(BaAn+Ch) ratios confirmed the pyrogenic origin for the PAHs accumulated in the dated sections between late 1960s and early 2000s, 10

but a predominant petrogenic origin afterwards (Fig. 6). The significant correlation between BaAn/Ch and BaAn/(BaAn+Ch) ratios and MS in C3 (P < 0.05; r = 0.71, r = 0.69, respectively) suggested the role of detrital material as carrier of these homologues; whereas in C4, where highest MAR were found, the negative correlations (P < 0.05; r = -0.82, r = -0.80) indicated that detrital material supply is

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diluting the homologue concentrations at this site. Despite evidences of hydrothermal activity in LC, no PAHs, n-alkanes or other aromatics have been detected in its tars and bitumen [58]. The PAHs accumulated in the sediments of LC are the result of a combination of pyrogenic and petrogenic sources. Petrogenic PAHs are originated from crude oil, fuels, lubricants and their derivatives; whereas the pyrogenic PAHs, usually released from industrial

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processes in the form of exhaust and solid residues, might reach the aquatic environment by direct

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atmospheric deposition or via contaminated soil runoff [59]. Low values of BaAn/(BaAn+Ch) ratios

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have been associated with LRAT of pyrogenic PAHs, because BaAn is more vulnerable to

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photodegradation than Ch during atmospheric transport [60]. Thus, the PAHs found in LC sediments

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could have resulted from short range and long range atmospheric transport from the nearby urban areas

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and from other industrial areas in the catchment. Direct discharges of untreated domestic and industrial wastes transported from the catchment through the Lerma River, the runoff from farms and farmlands

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surrounding the lake, and the exhausts and accidental oil spills from the fishing and tourist boats that

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frequent LC, might have also contribute to PAHs contamination.

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4.3.2. PCBs

In total, 209 PCBs were analyzed, either as individual or coeluting congeners, in the sediment samples but just 36 of them were detected, with chlorination levels ranging from 2 to 7: di-CB (PCB 8, 15), triCB (PCB 16, 17, 18, 25, 26, 28/31, 32), tetra-CB (PCB 41/47, 42, 43/52, 44, 47/48/62/65/75, 49, 56/60,

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61/63, 64/72, 66, 70, 74), penta-CB (PCB 87/111, 88/95, 99, 105, 110, 101/113, 118, 126), hexa-CB (PCB 132/153/168, 138/158, 139/149, 160/163/164) and hepta-CB (PCB 170/190, 180/193, 182/187). The ∑PCB concentrations in core C4 were generally lower than in core C3, with the maximum value at the core surface (Table 2). However, there was a west-to-east increase in ∑PCB concentrations

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in surface samples from core site C1 (9 ng g-1) to core C4 (27 ng g-1). These values in LC were generally higher than those published for remote and slightly contaminated urban sites such as Venice Lagoon, Italy [61], Lake Thun, Switzerland [62], Nepal [63] and Minnesota lakes, USA [64]; they were comparable to values reported for El Tule and Santa Elena lakes in the same region, the Central Mexican Plateau [7], but lower than in other sites near Mexico City [15, 18, 19] and in Tabasco [14].

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The ∑PCB concentrations recorded in LC were below the TEL benchmark of 34.1 ng g-1 [51]. Despite

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the low concentrations, PCB bioaccumulation in biota could potentially produce chronic effects, such

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as developmental effects, reproductive failure, liver damage, cancer, wasting syndrome, and death

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(NRC, [65]).

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Chlorination of the PCB congeners in core C3 varied between di-CB and penta-CB, and in core

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C4 between di-CB and hepta-CB (Fig. 4). Core 3 showed a clear increment in ∑PCB concentrations from the beginning of the 1990s towards the present, with the predominant congeners (>40% of the

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∑PCB) in those sections being tri-CB (PCB 18, 28/31) and tetra-CB (PCB 44, 49, 64/72) types, classified as non-dioxin-like PCBs [64]. Congeners at the C4 surface were predominantly (>36% of the

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∑PCB) dioxin-like penta-CB (PCB 105, 118) along with hexa-CB types (PCB 132/153/168, 138/158)

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that are often used as indicators of PCB presence [66]. In cores C3 and C4, ∑PCB fluxes (Fig. 5) were mostly constant from the middle of the 1960s

until a drastic increase at the core surface. These fluxes in cores C3 (0.2 - 4.3 ng cm-2 yr-1) and C4 (0.1 - 8.2 ng cm-2 yr-1) were comparable with values reported for two other lacustrine systems in the region, Santa Elena (0.5-6.0 ng cm-2 yr-1) and El Tule (0.2-3.3 ng cm-2 yr-1) where PCBs delivery is attributed 12

to LRAT [7]. None of the PCBs congeners correlated significantly with MS, supporting the hypothesis of atmospheric transportation and later deposition of PCBs in LC. From the beginning of the PCBs commercialization in 1929 in the USA and their introduction in the Mexican market in 1940s, PCBs were synthesized as multi-component mixtures with various

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grades of chlorination with the trade name Aroclor [67, 68]. The PCB-congener compositions of surface (C1 and C2) and core samples (C3 and C4) were compared by PCA with the congener blend of the most regularly used Aroclors in USA and Mexico (1016, 1221, 1232, 1242, 1248, 1254, 1260 and 1262). Principal components PC1 and PC2 explained 65.5% of the cumulative variance (Fig. 7). Highly chlorinated homologues were found in PC1, ranging from hexa-CB to nona-CB jointly with Aroclors

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1260 and 1262, whereas PC2 was composed mainly of lower mono-CB to penta-CB homologues. The

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PCB composition of C1 and C2 core surfaces, as well as some sections from sediment core C3 (i.e.

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sections 0 - 1, 7 - 8, 38 - 39, 84 - 85), approached the configuration of lighter Aroclors 1221, 1232 and

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1242, imported into Mexico from Monsanto Chemical Co., and primarily used as machine oil until

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1971; when they were internationally banned in 1974, Aroclor 1242 was modified to Aroclor 1016

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4.3.3. PBDEs

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[69], a configuration resembled by most of the C4 sections.

PBDE congeners were analyzed only in core C3 and surface samples C1 and C2. Those detected were

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lower-brominated types such as BDE 28 (2,4,4′-TriBDE), BDE 47 (2,2′,4,4′-TetraBDE), BDE 99

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(2,2′,4,4′,5-PentaBDE) and BDE 100 (2,2′,4,4′,6-PentaBDE). The prevalence of only lower brominated congeners, indicates that these compounds are most likely delivered by LRAT. The ΣPBDE concentrations in core C3 (0.2 - 2.5 ng g-1) were similar (Table 3) to the values observed in mountain lakes in Nepal [63] and Mexico [7] but lower than in other urban sites, such as the River Seine [2] or Lake Thun [62]. In general, core C3 had background PBDE values (< 2 ng g-1) from the middle of the 13

1960s to the beginning of the 2000s (Fig. 5); BDE 47 was the only congener detected in those sections and in the surface samples C1 and C2 (Table 2). However, the higher ΣPBDE concentration at the surface of core C3 incorporated the rest of the congeners detected, in which BDE 100 exceeded the safety limit (0.4 ng g-1) for sediments [70]. Thus, PBDE sediment concentrations in LC may represent a

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risk for aquatic biota, due to potential hormone disruption and reproductive failures [70]. The prevalence of congener BDE 47 in LC sediments might be related to the extensive use of commercial low-brominated mixtures Bromkal 70-5DE and Bromkal DE-71, as these penta-BDE mixtures were globally used as flame retardants in homes and workplaces (e.g. electric devices and paints) until 2004 [71] and are composed of BDE 28 (0.1 - 0.2%), BDE 47 (38 - 43%), BDE 99 (45 - 49%) and BDE 100

U

(8 - 13%) [72]. ΣPBDE fluxes in core C3 (Fig. 4) ranged between 0.02 to 0.40 ng g-1 yr-1 and were

A

[7]) and in Qinghai, China (0.58 ng g-1 yr-1, [73]).

N

comparable to values reported for other lakes in the Central Mexican Plateau (0.04 - 0.66 ng g-1 yr-1;

M

In summary, POP concentrations in LC might result from diverse sources, including direct discharges,

D

atmospheric deposition and soil runoff; and the temporal POP concentrations since 1960s are consistent

TE

with the urban and industrial growth of Guadalajara City metropolitan area; and most likely also reflects the agriculture and touristic development nearby and at longer distances of LC. Temperature is

EP

colder in the mountains bounding Chapala basin, where the POPs transported by LRAT would

CC

condense and deposit onto soils, which would be transported by winds and runoff to the lake. Both water-level decline and wind, promote clay resuspension in LC [74] which could enhance the

A

interaction and sorption of organic contaminants [75] delivered to the lake.

5. Conclusions 210

Pb-dated sediment records from Lake Chapala recorded the presence of POPs since ~50 years before

sampling (2010-2012) and increasing trends of POP contamination are in agreement with the urban and 14

industrial development of Guadalajara City metropolitan area. ΣPAHs, ΣPCBs and ΣPBDEs concentrations were comparable to moderate and highly contaminated lacustrine systems in Mexico and other sites in the world. Despite the restrictions on the use of PCBs (1971) and PBDEs (2004) the sediment records showed that PCB and PBDE concentrations and fluxes were still increasing, which

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suggest that these POPs are still used, or released to the environment owing to improper disposal. Regarding the origin of POPs in LC, (a) ratios of PAH homologues indicated a mixture of petrogenic and pyrogenic sources; (b) PCB-congener compositions approached the multi-component mixtures Aroclors 1221, 1232, 1242 and 1016 that were used in Mexico since 1940s; and (c) the prevalence of congener BDE 47 suggested that low-brominated mixtures Bromkal 70-5DE and Bromkal DE-71 were

U

the possible sources of PBDEs. Moreover, the domain of light-molecular-weight PAHs and preferential

N

presence of lower-chlorinated PCB congeners (di- to hepta-CB) and lower-brominated PBDEs (tri- and

A

penta-BDE) in the lacustrine sediments suggest that a portion of these LMW chemicals resulted from

M

long-range transport by winds from distant sources, including Guadalajara City or industrial areas

D

along the catchment. Finally, according to international benchmarks, PAH and PBDE sediment

TE

concentrations in LC are likely to cause biological effects and therefore the potential ecological risk posed by the sediments for biota and the human population due to fish consumption should not be

CC

EP

ignored and should continue to be considered in further studies.

Acknowledgments

A

This work was funded by the CONACYT grants SEMARNAT108093, INFR-2013-01/204818 and a PhD fellowship (JFOC); and the partial support from PROMEP/103.5/13/9335. Thanks are due to G. Ramírez-Reséndiz and C. Suárez-Gutiérrez for technical assistance, E. Cruz-Acevedo for figure edition, and three anonymous reviewers and Ms. Ann Grant for reviewing the English grammar and providing constructive comments on this manuscript. 15

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Captions Fig. 1. Location of sampling sites in Lake Chapala, Central Mexico. Land use and vegetation types

A

CC

EP

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D

M

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U

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were taken from INEGI [35].

25

Fig. 2. Depth profiles of activities and chronologies of

210

Pb (black dots) and

239,240

Pu (red dots) in

sediment cores from Lake Chapala, Mexico: a) C1; b) C2; c) C3 and d) C4. Dashed lines indicate the Pbsup fraction.

A

CC

EP

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210

26

Fig. 3. Grain size distribution (%: silt + clay), magnetic susceptibility (MS, K x 10-5 SI), organic matter (OM, %), δ13C (‰), δ15N (‰) and C/N ratio in sediment cores from Lake Chapala. Color code and the older dated section of each core: C1 (brown, 1915±2 years), C2 (green, 1926±17 years), C3 (black,

A

CC

EP

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D

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A

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1918±13 years) and C4 (blue, 1968±4 years).

27

Fig. 4. PAH, PCB and PBDE concentrations (solid lines; ng g-1) and fluxes (dashed lines; ng cm-2 yr-1)

A

CC

EP

TE

D

M

A

N

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in sediment cores C3 (black) and C4 (blue) from Lake Chapala, Mexico.

28

Fig. 5. Distributions of high-molecular-weight (HMW) and low-molecular-weight (LMW) PAH and

A

CC

EP

TE

D

M

A

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PCB homologues in sediment cores C3 and C4 from Lake Chapala, Mexico.

29

Figure 6. PAH ratios in sediment cores C3 (black) and C4 (blue) from Lake Chapala, Mexico. Type of combustion: An/(An + Phe) ratio: petroleum combustion <0.10; heavy fuel combustion >0.10; Flt/(Flt + Py) ratio: petrogenic <0.4; fossil fuel combustion (vehicle and crude oil) 0.4-0.5; coal + biomass

A

CC

EP

TE

D

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A

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emissions >0.5.

30

Fig. 7. PCA score plot of PCB homologue composition of sediment cores C3 and C4 from Lake

A

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TE

D

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Chapala in comparison with the composition of Aroclor standard mixtures.

31

Table 1. Location of sediment cores taken from Lake Chapala, Mexico. Localization

Water depth (m)

C1*

20°14′39″N, 103°15′50″W

7.5

Core length (cm) 108

C2*

20°18′41″N, 103°02′21″W

6.2

100

C3 C4

20°17′29″N, 102°58′26″W 20°14′57″N, 102°48′56″W

6.5 3.7

92 47

A

CC

EP

TE

D

M

A

N

U

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Core

32

A

N U SC RI PT

18.7 3.7 23.9 5.2 ND ND 4.9 11.1 ND 3.7 23.9 1.5ª 1.1ª 1.4ª 1.1ª 1.0a 0.7ª 0.4ª 1.1ª 0.6ª 0.5ª 1.2ª 1.5ª 0.5ª 0.9ª 0.7ª 1.0a 1.2ª 0.9ª 0.7ª 0.7ª 0.9ª 1.1ª 0.4ª 0.4 1.5 NA 3.5

Phe

An

19.3 151.0 33.9 10.4 79.8 12.3 32.6 19.4 9.8 8.9 59.5 28.9 4.6 49.5 10.1 8.2 15.8 5.0 5.3 17.2 6.3 6.8 17.5 8.5 ND 26.2 11.5 4.6 15.8 5.0 32.6 151.0 33.9 8.9 16.8 5.9 7.0 12.3 2.5 8.1 14.2 2.8 7.3 12.1 2.8 5.1 10.4 2.3 3.4 7.0 1.5ª 2.8 6.3 1.2ª 3.6 8.3 1.6ª 2.5 6.9 1.3ª 2.5 7.2 1.2ª 2.4 6.6 1.3ª 2.1 6.4 1.4ª 1.8 5.2 1.1ª 1.4ª 4.4 0.8ª 1.8 9.3 3.3 2.1 5.6 1.1ª 2.6 6.7 1.3ª 2.0 5.9 0.7ª 3.4 4.9 0.9ª 3.9 7.2 1.2ª 2.7 5.9 1.2ª 3.0 5.4 0.9ª 2.5 5.2 1.0a 1.4 4.4 0.7 8.9 16.8 5.9 15.4 85.5 17.1 14.2 34.6 7.1

Flt

Py

235 899.0 24.1 39.2 11.3 12.3 25.7 33.2 21.4 25.4 5.7 7.5 7.6 6.6 6.5 6.5 11.4 11.8 5.7 6.5 235 899 19.0 21.8 9.6 15.0 10.3 18.4 8.6 12.2 6.9 10.0 4.9 8.0 4.3 6.9 5.6 8.7 4.1 7.6 4.1 8.6 3.3 7.6 3.6 7.3 4.5 6.7 2.6 4.3 7.9 9.9 2.7ª 4.4 3.7 5.4 3.4 5.0 2.6 3.8 3.5 4.5 3.1ª 3.2ª 2.6ª 3.2ª 3.0a 3.7 2.6 3.2ª 19.0 21.8 49.8 93.9 16.1 25.2

M

10.2 10.9 12.1 9.0 3.7 9.7 4.8 7.5 20.1 3.7 20.1 6.4 2.4 2.7 2.8 2.2 1.7a 1.5a 1.6ª 1.6ª 1.9ª 1.4ª 1.8ª 1.2ª 1.1ª 2.8 1.3ª 2.1 1.3ª 1.2ª 1.5ª 0.8ª 0.9ª 0.9ª 0.8 6.4 6.0 5.8

Fl

ED

13.0 50.7 11.4 48.1 11.7 39.2 8.9 30.0 62.5 8.9 62.5 0.5 63.0 2.5 40.9 4.5 57.0 6.5 47.6 8.5 32.7 10.5 24.6 12.5 18.2 14.5 25.1 16.5 18.4 18.5 18.9 20.5 16.5 22.5 24.3 24.5 18.1 26.5 13.6 28.5 9.6 30.5 20.8 32.5 25.9 34.5 19.8 38.5 17.9 40.5 29.3 42.5 10.0 44.5 18.5 46.5 14.8 9.6 63.0 Surface 12.3 Surface 12.4

Acy Ace

CC E

C3 C3 C3 C3 C3 C3 C3 C3 C3 Min Max C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 C4 Min Max C1 C2

Na

PT

Sample

High-molecular-weight PAHs BaAn

Ch

8.9 1.0a 0.4ª 1.1ª 0.8ª 0.8ª 0.7ª 0.7ª ND 0.4ª 8.9 15.4 2.5 2.9 2.6 1.9 6.1 5.0 6.5 5.6 5.8 4.9 4.7 3.6 3.1 11.6 3.6 3.8 3.2 3.2 3.8 4.3 3.2 3.5 1.9 15.4 1.2 0.9

82.6 5.3 2.5 9.4 4.7 3.5 4.2 2.9 ND 2.5 82.6 23.4 7.2 9.4 6.3 5.4 6.2 5.1 6.6 5.7 5.9 5.0 4.8 3.6 3.2 11.9 3.6 3.9 3.3 3.3 3.9 4.4 3.3 3.6 3.2 23.4 4.3 6.0

A

Low-molecular-weight PAHs Depth (cm) 0.5 7.5 15.5 22.5 30.5 38.5 45.5 69.5 84.5

B(b + k)Flt BaPy 4.4 4.0 1.5ª 7.9 5.2 ND 3.2 1.6ª ND 1.5ª 7.9 47.0 6.5 6.2 6.7 5.9 8.7 0.5ª 9.4 6.3 7.5 6.7 5.1 4.9 4.0 58.3 5.6 5.7 5.3 4.7 5.5 8.1 5.5 5.8 0.5 58.3 4.6 4.9

5.7 33.9 4.6 38.2 27.2 20.4 25.7 15.6 26.9 4.6 38.2 16.9 2.3ª 2.1ª 2.5ª 2.1ª 1.6ª 1.2ª 1.3ª 1.1ª 1.3ª 1.1ª 1.4ª 1.2ª 0.7ª 10.3 0.8ª 0.7ª 0.5ª 0.4ª 0.5ª 0.8ª 0.4ª 0.5ª 0.4 16.9 7.9 19.8

IPy

ND ND ND ND ND ND ND ND ND 22.4 3.2 2.8 3.6 3.4 2.5ª 2.3ª 0.1ª 2.5ª 2.4ª 2.5ª 2.4ª 1.9ª 1.7ª 15.1 1.4ª 1.5ª 1.5ª 1.0a 1.5ª 2.4ª 0.9ª 1.5ª 0.1 22.4 ND ND

DahAn BghiPe ND ND ND ND ND ND ND ND ND 5.4 1.2ª 0.8ª 1.7ª 1.1ª 0.2ª 0.1ª 0.4ª 0.2ª 0.1ª 0.5ª 0.5ª 1.8 0.2ª 3.4 ND ND 0.3ª 0.2ª 0.5ª 1.1ª 1.0a 0.2ª 0.1 5.4 ND ND

ND ND ND ND ND ND ND ND ND 23.3 4.4 4.1 4.5 3.9 3.8 3.2 3.4 3.5 4.0 3.5 3.2 2.5 1.9ª 14.9 1.9ª 2.1ª 1.7ª 1.3ª 1.5ª 2.1ª 0.2ª 0.7ª 0.2 23.3 110 85.6

Total POP concentration Pe

BePy

ƩPAHs

ƩPCBs

ƩPBDEs

47.2 105 13.9 144 141 90.6 161 134 156 13.9 161 327 380 359 391 306 287 298 412 295 396 345 318 305 319 390 377 415 375 361 442 608 421 438 287 608 1.1ª 4.6

3.3 12.5 2.2 10.5 2.8 ND 2.3 ND ND 2.2 12.5 24.4 4.8 4.7 5.2 4.1 3.4 2.9 3.3 3.3 3.3 3.1 2.9 2.7 1.9 15.1 1.8 2.1 0.1ª 0.4ª 1.7ª 2.2 1.3ª 1.6ª 0.1 24.4 2.6 4.9

1482.3 275.2 141.8 275.1 164.4 115.7 95.4 115.2 170.6 95.4 1482.3 648.6 503.1 507.4 518.7 404.4 370.9 360.1 498.4 366.3 471.1 412.4 391.2 366.2 364.8 576.1 434.5 484.0 430.0 410.8 512.3 660.9 471.9 487.2 360.1 660.9 411.8 245.6

27.1 12.3 7.5 5.5 3.0 5.2 3.8 6.1 5.1 3.0 27.1 24.3 0.8 1.5 2.0 2.3 0.7 0.5 0.4 4.2 1.3 6.2 1.2 0.5 ND ND 0.9 0.6 0.4 0.9 2.0 0.9 0.6 0.3 0.3 24.3 9.0 11.3

2.5 0.4 1.1 0.4 0.2 0.5 NA NA NA 0.2 2.5 NA NA NA NA NA NA NA NA NA NA NA NA NA NA NA NA NA NA NA NA NA NA NA 0.6 0.4

33

2.1 2.0 3.5 1.8 34.6 5.87 6.71 21.2 NA NA NA 190

1.7 41.9 560

1.9 46.9 220

3.2 111 750

N U SC RI PT

Detection limit TELa LELa

3.6 53 490

1.9 31.7 320

2.2 57.1 340

2.0 NA 240c

3.1 31.9 370

2.6 NA 200

1.3 6.22 60

2.4 NA 170

4.4 NA NA

1.9 NA NA

NA 4000

0.2 34.1 70

0.2 NA NA

Table 2. PAH, PCB and PBDE concentrations (ng g-1) in sediment cores from Lake Chapala, Jalisco, Mexico.

a

Analyte concentration below the method detection limit. ND, not detected. NA, not available. Buchman, 2008 c The value corresponds to B(k)Flt.

A

CC E

PT

ED

M

A

b

34

N U SC RI PT

Table 3. PAH, PCB and PBDE concentrations (ng g-1, dry weight) in freshwater sediments in Mexico and elsewhere. Type of environment

∑PAHs

∑PCBs

∑PBDEs

Reference

Mecoacán Lagoon, Tabasco

Urban, contaminated

0.1 - 35.7

6 - 372

NA

[11]

Chalco Lake, Mexico City

Urban, contaminated

287

621

NA

[16]

Lake Texcoco, Mexico City

Urban, slightly contaminated

95

64

NA

[16]

Espejo de los Lirios Lake, Mexico City

Urban, slightly contaminated

122

253

NA

[17]

Urban, slightly contaminated

17.6

NA

NA

[14]

Urban, contaminated

259 - 1176

23.8 - 77

NA

[12]

Rural slightly contaminated

50.4 - 853

1.7 - 24.7

0.3–1.5

[15]

Santa Elena Lake, Jalisco

Rural, slightly contaminated

119 - 482

1.5 - 15.4

0.4–1.8

[15]

Lake Chapala, Jalisco-Michoacán

Urban, slightly contaminated

95.4 - 1482

0.3 - 27.1

0.2 - 2.5

**

Venice Lagoon, Italy

Urban, slightly contaminated

NA

0.6 - 18.4

NA

[47]

Ny-Ålesund lakes, Norway

Remote, slightly contaminated

11 - 1100

0.06 - 21

0.024 - 0.97

[37]

Lake Thun, Switzerland

Urban, slightly contaminated

NA

0.4 - 3.8

0.1 - 5.1

[48]

Location

Las Matas Lagoon, Veracruz

ED

El Tule Lake, Jalisco

M

Tecocomulco Lake, Hidalgo

A

Mexico

CC E

PT

Other sites around the world

Himalayan lakes, Nepal

Remote, slightly contaminated

67.9 ± 22.1

1.00 ± 0.54

1.16 ± 0.71

[49]

Lake Taihu, China

Urban, contaminated

239 - 1167

1.08 - 2.79

NA

[38]

Seine River basin, France

Urban, contaminated

5000 - 90,000

500 - 2371

15 - 53

[2]

Minnesota lakes, USA

Urban and rural, contaminated

489.3 ± 980.5

18.8 - 19.7

NA

[50]

A

NA= not available.

**The present study

35