Applied Geochemistry 41 (2014) 34–48

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Applied Geochemistry

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Deciphering the impact of land-uses on terrestrial organic matterand mercury inputs to large boreal lakes of central Québec using ligninbiomarkers

0883-2927/$ - see front matter � 2013 Elsevier Ltd. All rights reserved.http://dx.doi.org/10.1016/j.apgeochem.2013.11.008

⇑ Corresponding author. Tel.: +1 514 987 3000x3523; fax: +1 514 987 3635.E-mail address: matthieumoingt@yahoo.fr (M. Moingt).

Matthieu Moingt ⇑, Marc Lucotte, Serge Paquet, Bassam GhalebGEOTOP, Institut des Sciences de l’Environnement, Université du Québec à Montréal, C.P.8888, suc. Centre-Ville, Montréal, Québec H3C 3P8, Canada

a r t i c l e i n f o

Article history:Received 17 June 2013Accepted 17 November 2013Available online 1 December 2013Editorial handling by JoAnn M. Holloway

a b s t r a c t

To evaluate watershed impacts of anthropogenic activities on terrestrial organic matter (TOM) and totalmercury (THg) dynamics in large boreal lake ecosystems, we studied sediment cores retrieved in eightlarge lakes of Québec (Canada). Two lakes with pristine watersheds were considered as reference lakesand six lakes with watersheds affected by different types of anthropogenic activities (e.g. logging and/or mining activities) were used to illustrate the influence of land-use on TOM and Hg cycling in lakes.A Geographical Information System (GIS) approach was used to correlate the evolution of anthropogenicland-uses from 1979 to 2010 (e.g. logging and mining activities) to TOM and THg contents measured insediment cores. In each core, THg concentrations gradually increased over the recent years. Using ligninbiomarkers, we noticed that the presence of both intense logging and mining activities in the watersheddoes not necessarily correspond to noticeable changes in the relative amount of terrestrial organic matter(TOM) exported from the watershed to the sediments and by extension to the level of THg measured insediments. Apparently large-scale watersheds show some ‘‘buffering’’ capacity to land-use disturbance.

� 2013 Elsevier Ltd. All rights reserved.

1. Introduction

Mercury (Hg) dynamics in lakes in central Québec are of specialconcern as local populations frequently consume wild fish and arethus chronically exposed to the contaminant. For over a century,the release of Hg to the atmosphere associated with anthropogenicactivities resulted in an increase in the heavy metal deposition inboreal lakes ecosystems (Pacyna et al., 2006). Dry and wet Hgdeposition from anthropogenic sources can reach areas verydistant from their point source emission (Fitzgerald et al., 1998;Pacyna et al., 2006), thus remote areas such as central Québeccan then be impacted. Hg is then transferred to the lake andaccumulated in sediments where increases in total mercury(THg) contents have been reported (Lucotte et al., 1995; Rognerudand Fjeld, 2001). In parallel, central Quebec watersheds arefrequently perturbed by multiple anthropogenic activities such aslogging, mining, road construction and urbanization. Those pertur-bations modify the natural Hg fluxes from the watersheds to theaquatic systems and could be responsible for supplementaryamounts of Hg reaching the lakes, which have the potential to beincorporated into the food web (Ikingura and Akagi, 1996; Lavoieet al., 2010; Soto et al., 2011; Watras et al., 1998).

It is recognized that terrestrial organic matter (TOM) plays amajor role in the transport of contaminants from the watershedto the lake (Coquery and Welbourn, 1995; Moingt et al., 2010;Ouellet et al., 2009; Teisserenc et al., 2010). Lake sediments repre-sent a powerful tool to reconstruct both Hg and TOM transfersfrom the watersheds to the aquatic systems. Lignin biomarkershave been used to evaluate both TOM fluxes and quality reachingsediments (Dittmar and Lara, 2001; Hedges et al., 1997; Hu et al.,1999; Ouellet et al., 2009) and lignin degradation biomarkers havebeen widely exploited for the study of TOM dynamics in soils,water column, and sediments (Chabbi and Rumpel, 2004; Houelet al., 2006; Ouellet et al., 2009). In comparison with bulk carbonanalysis, lignin biomarkers in sediment cores bring new informa-tion on the nature of TOM fluxes from watersheds to lakes and helpto reconstruct historical watershed modifications. Moreover,because Hg is known to have affinity for TOM (Kainz et al., 2003;Kolka et al., 1999; Yang et al., 2007), lignin biomarkers could yieldnew information on Hg dynamics in boreal lakes. Lignin alkalineoxidation by copper oxide (CuO) produces short-chain phenolscomprising ketone, aldehyde and/or carboxylic acid functionalitiespresent in the parent lignin macropolymer. These major oxidationproducts are grouped into four major monomer families: vanillyl(V), syringyl (S), cinnamyl (C) and p-hydroxyphenols (P) (Goñiand Hedges, 1992; Goñi et al., 1993; Hedges and Ertel, 1982).The relative proportions of these monomeric subunits allow dis-crimination between different types of vascular plant tissues, and

M. Moingt et al. / Applied Geochemistry 41 (2014) 34–48 35

have been used to decipher sources and relative degradation stateof organic matter in several natural ecosystems (Dittmar and Lara,2001; Hedges and Mann, 1979; Hernes and Benner, 2003). More-over, several studies have shown that lignin quality is influencedby hydrologic and soil processes in watersheds (Dalzell et al.,2007; Hernes et al., 2007).

In this paper, we combined the geochemical study of sedimentdata (210Pb dating, THg concentrations, lignin biomarkers) withwatershed characteristics identified through GIS (land-use andvegetation cover) to evaluate the impact over three decades ofanthropogenic watershed perturbations relative to the increase inHg transfers from watersheds to sediments in large lakes ecosys-tems. We sampled sediment cores in eight large boreal lakes withsurface areas ranging from 15 to 210 km2, six of them presentingspecific disturbances in their watershed such as intensive loggingor mining, and two unperturbed lakes chosen as reference lakes.Several studies point to anthropogenic perturbations occurring inthe watershed (e.g. logging and mining activities) as a contributorto the increasing amount of THg reaching lakes sediments. How-ever, these studies have focused solely on small lakes ecosystemswith surface areas ranging from 0.4 to 1.3 km2 (Garcia and Cari-gnan, 2000, 2005; Porvari and Verta, 2003; Porvari et al., 2003).Large lake ecosystems have been ignored despite the fact that theyrepresent a better picture of socio-economic-environmental land-uses of the boreal forest region (fishery, mining and logging activ-ities). Moreover, small lakes only allow the study of one processone at a time whereas large lake ecosystems integrate a sum ofvarious processes and include the inertia of the system. Such ap-proaches will help to gain a more holistic understanding of the cy-cling of Hg between terrestrial and aquatic systems and theinfluence of anthropogenic effects and environmental changes inthe transport of Hg into aquatic ecosystems.

2. Material and methods

2.1. Study sites

Matagami, Ouescapis, Rodayer, Waswanipi, Dickson andChibougamau lakes are located in the administrative region ofNorthern Quebec (Fig. 1). Lakes characteristics (location, altitude,lake area, drainage area, ratio of drainage area to lake area (DA/LA)) are specified in Moingt et al. (2013). Matagami Lake is a43 km long and 14 km wide lake in a swampy region of NorthernQuebec and represents the confluence point of Allard, Bell, Goua-ult and Waswanipi rivers. Perturbations in the watershed includeroad construction, logging, mining and golf activities. Ouescapisand Rodayer lakes are located approximately 70 and 120 kmnorth of Matagami Lake, respectively. None of these lakes showany substantial perturbations in their watershed. Waswanipi Lakeis situated about 100 km south-east from Matagami Lake towhich it is linked upstream by the Waswanipi River. WaswanipiLake presents a watershed perturbed with logging and miningactivities and several gravel roads. Dickson Lake is located about180 km south-east from Matagami Lake. This lake has a history ofintense logging activity in its watershed. Chibougamau Lake is lo-cated approximately 10 km south of the town of the same name.This lake is the source of the Chibougamau River and is inter-spersed with several islands, deep and large bays. The watershedof Chibougamau Lake has been intensely used for mining activi-ties (Petit et al., 2011). Des Jardins East and West lakes (total areaof 15.02 km2) are located in the Témiscamingue region furthersouth (Fig. 1). Des Jardins East Lake flows into Des Jardins WestLake and their watersheds are covered by mature mixed woodforests. Perturbations in the watershed include gravel roads andlogging activities.

2.2. Landscape and historical analysis

Landscape analyses were performed following the method de-scribed in details by Moingt et al. (2013). Briefly, landscape charac-terization was performed with raster satellite images from Landsat7 satellite imagery with image processing using GIS GeographicResources Analysis Support System (GRASS 6.4; http://grass.-fbk.eu/) software in order to decipher the morphology and the landcover of watersheds. The Canadian digital elevation data (scale: 1/50,000), extracted from the hypsographic and hydrographic ele-ments of the National Topographic Data Base (Tahvanainen andHaraguchi, 2013) were also used. For the purposes of our study,the watershed is defined as the land area draining either directlyinto a lake or into the first lake upstream corresponding to an order3 watershed level which has been described as the level influenc-ing the most the associated aquatic ecosystem (Beaulne et al.,2012; Slauenwhite and Wangersky, 1996). In order to determinethe mean watershed slope and to identify different slope classes(intervals of 2% from 0% to 20%, e.g. 0–2%) within each watershed,the slope of each pixel was calculated. Land cover was estimatedby applying the Maximum Likelihood Classification (MLC) algo-rithm modified by Neteler and Mitasova (2008) to the entire setof visible spectral bands. Historical reconstructions were basedon observations coming from satellite images over a 30 year period(1979–2011).

2.3. Sediment sampling

We used a pneumatic Mackereth corer to sample lake sedi-ments during sample missions in 2009 and 2010. This type of corerpresents the advantage of reducing both perturbations at thewater–sediment interface and sediment compaction. During thesampling, a 1.5-m long Plexiglas tube (diameter 10 cm) is slowlypushed into the sediment using compressed air. Then, sedimentcores were sub-sampled in 20 mL glass vials (pre-combusted at500 �C for 3 h to avoid carbon contamination of the vessels andcapped with Teflon� liners) every centimeter using a Teflon� spat-ula. For each slice we only conserved the center of the core to avoidcross-contamination between samples. Finally, samples werefreeze-dried prior to analysis. In order to integrate spatial varia-tions in watershed to lake TOM and Hg transfers, coring was per-formed at the focal point of the lake (Hakanson and Jansson,1983; Teisserenc et al., 2010).

2.4. Chemical analyses

2.4.1. 210Pb radiometric measurementsSedimentation rates were determined by radiometric measure-

ment of 210Pb activity using alpha spectrometric measurement ofthe activity of the daughter product 210Po (Flynn, 1968) andassuming secular radioactive equilibrium between the two iso-topes. Aliquots between 0.2 and 0.4 g of sediments were spikedwith 209Po yield tracer and digested in Teflon vials using a concen-trated 5:4:1 mixture of HNO3:HCl:HF. Addition of H2O2 was per-formed to eliminate organic carbon (OC) in the sample. Theresidue was then converted to a chloride salt by repeated evapora-tion with 6 M HCl. Before the last evaporation was complete, theremaining solution was centrifuged and the supernatant wascollected in order to get rid of black carbon. Then, approximately100 mL of 0.5 M HCl was added to the collected supernatant to de-crease the concentration of the solution. Then 1.5 mL of 1 M ascor-bic acid was added to reduce Fe3+ into Fe2+. Po isotopes wheredeposited on a spinning Ag disk (Hamilton and Smith, 1986) andthe activity measured by a-spectrometry using ORTEC siliconsurface barrier detectors coupled with a PC running under Maestrodata acquisition software. Blanks were run through the same

Fig. 1. Location of studied lakes on a Quebec map.

36 M. Moingt et al. / Applied Geochemistry 41 (2014) 34–48

procedure. Chronology was established using downcore unsup-ported 210Pb, a constant initial concentration model (CIC), anddividing nonlinear profiles into linear segments assuming constantsedimentation for each of these segments (Appleby and Oldfield,1992). 210Pbex corresponds to the excess of 210Pb in sediments(i.e. the difference between the total 210Pb activity at depth (i)minus the value of 210Pb supported activity). 210Pb supported activ-ity corresponds to the average of measured values at the bottom ofthe core when 210Pb activity does not vary anymore.

2.4.2. Total mercury analysesReplicate THg analyses were performed by cold vapor atomic

fluorescence spectrometry (CV-AFS) following the protocol devel-oped by Bloom and Fitzgerald (1988) and adapted by Pichet et al.(1999). Briefly, a combination of 16 M HNO3:6 M HCl (10 mL:1 mL)is added to approximately 250 mg of ground, freeze-dried sedi-ment and then heated to 120 �C for 6 h. The remaining solutionis brought back to a volume of 30 mL with NANOpure� waterand analyzed by atomic fluorescence. Hg calibration was done byinjecting known quantities of Hg (II) (400–1000 pg of Hg). Thedetection limit for a 250 mg sample was 0.1 ng/g. The accuracyof the method was verified using the Mess-3 certified standard(NRC Canada). With an average value of 87 ± 3 ng/g for seven ali-quots, our results fell within the certified value (92 ± 9 ng/g).

2.4.3. Lignin biomarkers analysesIn order to conduct lignin biomarkers analyses, percentage of

OC for each sample was measured performing total carbon and

nitrogen analyses using a CE-Instruments model NC2500™ ele-mental analyzer with a relative precision of ±5% (1r). Replicatemeasurements on samples, treated and untreated by vapor acidifi-cation, were performed to verify that the inorganic carbon fractionwas negligible. Acidified sample measurements systematically fellwithin the range of the non-acidified sample, thus demonstratingthat total carbon content can be considered as OC content.

Sediment samples were subjected to alkaline cupric oxide (CuO)oxidation according to a modified protocol from the initial methoddescribed by Goñi and Montgomery (2000). For each sample,3.0 ± 0.1 mg of OC and CuO (330 ± 4 mg) were mixed in about3.2 mL of 2 N NaOH in a reaction bomb purged with N2. Oxidationswere conducted with a Hewlett Packard 5890A™ gas chromato-graph oven modified by PRIME FOCUS Inc. (Seattle, WA) and fittedwith 12 Monel-400 alloy (Ni:Cu:Fe 65:32:3 wt%) reaction vessels.Reaction bombs were heated to 150 �C for 150 min with an initialtemperature gradient of 4.2 �C/min for 30 min. After oxidation,50 lL of the internal standard cinnamic acid and ethyl-vanillinwas added to the bomb and the supernatant was decanted andacidified to pH 1 with HCl (2 N). The organic phase was then li-quid–liquid extracted using ethyl acetate and dried by rotary evap-oration. Finally, the extract was dissolved another time in pyridineand derivatized with N,O-bis(trimethylsilyl)trifluoroacetamide(BSTFA) and trimethylchlorosilane (TCMS; 99:1). A 2-lL extractionfraction was analyzed by GC/MS (Varian 3800/Saturn 2000™)fitted with a fused capillary column (DB-1 from J&W, 60 m,320 lm). The injector and the detector were held at 300 �C andHe was used as carrier gas in splitless mode. The initial column

Table 1Comparison and variability of lignin CuO oxidation parameters measured from a reference sediment: SAG 05 (estuarine sediment). Values from Louchouarn et al. (2000) andHouel et al. (2006) were generated using conventional oven and microwave oven respectively.

SAG 05

This study (n = 9) Louchouarn et al. (2000) (n = 11) Houel et al. (2006) (n = 6)

Average SD Average SD Average SD

Lambda 3.50 0.23 3.53 0.17 2.81 0.31Sigma 8 7.41 0.48 ND ND ND NDS 0.70 0.20 0.56 0.05 0.44 0.03V 2.57 0.29 2.73 0.11 2.11 0.29C 0.23 0.04 0.23 0.04 0.26 0.04P 0.30 0.03 0.54 0.10 0.34 0.04S/V 0.28 0.11 0.21 0.02 0.21 0.03C/V 0.09 0.02 0.08 0.01 0.13 0.02P/(V + S) 0.09 0.01 ND ND 0.14 0.023,5-Bd/V 0.10 0.02 0.08 0.02 0.11 0.01(Ac/Ad)S 0.20 0.07 0.38 0.05 0.47 0.13(Ac/Ad)V 0.37 0.03 0.38 0.05 0.48 0.09

M. Moingt et al. / Applied Geochemistry 41 (2014) 34–48 37

oven temperature was set at 100 �C with a temperature gradient of4 �C/min to 320 �C followed by a holding time of 10 min.

Replicate analyses of the SAG 05 ‘‘standard’’ sediment sample(n = 9) showed that the analytical variability of the alkaline CuOoxidation products and major indicators range from 6% to 38% (Ta-ble 1). The intercomparison of this standard sediment analysesshows that the composition parameters and the yields of the pres-ent study are comparable to those obtained by both microwavedigestion and traditional oven method (Table 1).

2.4.4. Statistical analysesStatistical tests were applied to the regression observed

between the enrichment in THg from the deeper sections of thecores to the surface ones. Such tests were also applied to the mor-phological characteristics of the watersheds. In all cases, normaldistribution of residuals was tested by performing a Shapiro–Wilkgoodness of fit test, using a statistical analysis program (JMP 7) andaccepted W values >0.05. The F-ratio was then computed in orderto evaluate the effectiveness of the model and the student para-metric test (t-test) was processed to evaluate the pertinence ofregression parameters (p < 0.05).

3. Results

3.1. 210Pb dating

Figs. 2–4 present the results for radiometric measurementsrealized on lake sediment cores. In surface sediments, 210Pbactivities are 32, 42, 18, 37, 14 and 15 dpm/g for Ouescapis, Roda-yer, Waswanipi, Chibougamau, Dickson and Matagami lakesrespectively.

Water content of each sample was determined by weighingsediment before and after freeze-drying in order to correct the sed-iment accumulation rate because of possible compaction of thecore during sampling. After calculation, sedimentation rates ob-tained taking or not into account dry bulk density of each sedimentlayer were not different suggesting a minor impact of compactionin our case. Thus, sedimentation rates were directly estimatedusing the slope of the relationship between depth (cm), the naturallogarithm of 210Pbex and the 210Pb decay constant.

Desjardins Lake has an estimated sedimentation rate of0.22 ± 00.6 (Lucotte et al., 1995; Teisserenc et al., 2010). Estimatedsedimentation rates are 0.11, 0.05, 0.05, 0.03, 0.05 and 0.31 cm/year for Ouescapis, Rodayer, Waswanipi, Chibougamau, Dicksonand Matagami lakes respectively. Only Matagami Lake seems topresent a modification of its sedimentation rate through time

(Fig. 4). Indeed, we can observe an inflexion point around 6 cmdepth where the sedimentation rate rapidly shifts from 0.11 to0.31 cm/year. Due to the unusual 210Pb profile in Matagami Lake’ssediment core, 137Cs measurements have been performed in orderto verify the estimated sedimentation rates in this lake. Fig. 4 pre-sents 137Cs activity and 210Pb age previously estimated vs. depth inthe sediment core. 137Cs is an anthropogenic radionuclide whichhas been released in important amounts because of nuclear tests.137Cs started to fallout on land surfaces circa 1952 (Davis et al.,1984) and peaks of 137Cs are generally observed circa 1963 in sed-iments cores (Ali et al., 2008). In Fig. 4, we notice that the peak of137Cs corresponds to a 210Pb age of approximately 45 years whichis in agreement with the date of 1963. Thus, 137Cs confirms theages previously estimated by 210Pb dating.

Based on sedimentation rates estimated by 210Pb dating, astraight horizontal line symbolizes an age of 100 years in each sed-iment core in Figs. 5, 6, 10 and A1 and A2 from the Appendix A.

3.2. Spatial, historical analysis of lakes and drainage areas and THgcontents in lakes sediment cores

Spatial and historical analysis of lakes, drainage areas and THgcontents in lakes sediment cores of the present study have been pre-viously performed, measured and described by Moingt et al. (2013).Data from Moingt et al. (2013) summarized the principal bio-mor-phological parameters and THg results of the studied lakes are pre-sented respectively in Tables A1 and A2 from the Appendix A.

Based on 210Pb dating, the increases in THg in reference lakescorrespond to about the start of the industrial era (Appendix A,Fig. A1). The Hg anthropogenic sedimentary enrichment factor(ASEF) values come to 2.7 and 2.8 for Ouescapis and Rodayer lakesrespectively. These ASEF values are similar to those frequently re-ported in the literature for pristine North American lakes of theboreal domain (Lindberg et al., 2007; Lucotte et al., 1995).

Three out of six lakes (Des Jardins East, Des Jardins West andWaswanipi) with perturbed watersheds are characterized by HgASEF values inferior to 3 since the beginning of the industrial era.i.e. similar to those of the two pristine lakes (Appendix A,Fig. A2). The three remaining lakes (Chibougamau, Matagami andDickson) present much higher Hg ASEF values (3.8, 5.4 and 15respectively; Appendix A, Table A2).

3.3. Lignin biomarkers measurements

Lambda and Sigma 8 indicators are commonly used to estimatethe relative amount of TOM in sediments (Houel et al., 2006;

Fig. 2. Total 210Pb activity (full black rhombs) and natural logarithm of unsupported 210Pb activity (white rhombs) vs. depth in reference lakes sediment cores (i.e. Ouescapisand Rodayer lakes).

38 M. Moingt et al. / Applied Geochemistry 41 (2014) 34–48

Teisserenc et al., 2011). Lambda corresponds to the sum of vanillyl,syringyl and cinnamyl phenols per 100 mg of OC whereas Sigma 8represents the same sum of compounds but normalized for 10 g ofsample. Thus, Lambda corresponds to the relative amount of terrig-enous OM with respect to the total OC content of a sediment sam-ple and Sigma 8 corresponds to the relative amount of terrigenousOM to a sediment sample. Fig. 5 presents Lambda and Sigma 8indicators profiles for reference lakes (i.e. Ouescapis and Rodayerlakes) in the first 40 cm of the sediment cores. Lambda profilesare similar for Ouescapis and Rodayer lakes with values rangingfrom 0.71 to 1.51 mg/100 mg of OC and from 0.68 to 1.76 mg/100 mg of OC respectively. Sigma 8 ranges from 1.71 to 3.76 mg/10 g of sample and from 1.17 to 4.08 mg/10 g of sample for Oues-capis and Rodayer lakes respectively.

Fig. 6 presents Lambda and Sigma 8 indicators profiles in lakespresenting various types of perturbations in their watersheds. Con-sidering Lambda values, only Matagami Lake shows major devia-tion from the 0.5 to 2.0 mg/100 mg of OC range observed in thetwo reference lakes with values as high as 7.94 mg/100 mg of OC.On the other hand, Sigma 8 values present a wide range of valuescomprised between 0.02 and 48.14 mg/10 g of sample (Fig. 6).

In Figs. 7–9, we present the mean values of lignin biomarkersand the corresponding standard deviation for the last 100 year per-iod, estimated using 210Pb dating, of each sediment core. MeanLambda values are comprised between 0.77 ± 0.07 and3.59 ± 2.10 mg/100 mg of OC (Fig. 7) which is in agreement withvalues reported for large lake sediments in the literature (Hyodoet al., 2008; Orem et al., 1997; Petit et al., 2011; Tareq et al.,2011). Reference lakes present the lowest Lambda values. With

the exception of Matagami Lake, which present important Lambdavariations through time, all the lakes have Lambda values that arerelatively stable during the last 100 year period (Fig. 7).

S/V ratios are generally used to estimate the contribution ofangiosperm species to TOM contents in lake sediments (Hedgesand Mann, 1979; Tesi et al., 2008). Mean S/V ratio values vary from0.22 ± 0.07 to 0.39 ± 0.02 (Fig. 8). P/(V + S) ratio is used as indicatorof the degradation state of TOM (Dittmar and Lara, 2001; Opsahland Benner, 1995) since demethoxylation leads to the selectiveloss of OCH3 group in S and V families whereas p-hydroxyphenolsare not affected (Hedges and Ertel, 1982). 3,5-Bd/V is an indicatorof OM maturation in soils (Houel et al., 2006; Prahl et al., 1994).Mean P/(V + S) values range from 0.16 ± 0.05 to 1.07 ± 0.08whereas mean 3,5-Bd/V values are comprised between0.14 ± 0.06 and 0.70 ± 0.15. Fig. 9 shows that when all lakes areconsidered together, P/(V + S) and 3,5-Bd/V ratios are not corre-lated (r = 0.53; p = 0.2825). If Chibougamau Lake is excluded, wethen find a strong correlation between those two ratios (r = 0.95;p = 0.0150).

Lignin phenol results are summarized in Table A3 in the Appen-dix A.

4. Discussion

Although sampling one core per site at the focal point of a givenlarge lake does not allow observing site-specific inhomogeneities,it remains an efficient way to obtain a global signal integratingboth TOM and Hg dynamics at the watershed scale (Hakansonand Jansson, 1983; Moingt et al., 2013; Teisserenc et al., 2010).

Fig. 3. Total 210Pb activity (full black rhombs) and natural logarithm of unsupported 210Pb activity (white rhombs) vs. depth in Dickson, Chibougamau and Waswanipi lakessediment cores.

M. Moingt et al. / Applied Geochemistry 41 (2014) 34–48 39

4.1. Lambda and Sigma 8 indicators

In reference lakes, the relative stability of Lambda along thesediment cores (Fig. 5) indicates that little changes occurred inthe ratio of allochthonous vs. autochthonous organic matter overtime. The magnitude of variation in Sigma 8 along the core profilesis higher than that of Lambda (Fig. 5). Given that these two indica-tors have the same trends along the profiles, their variations areunlikely due to variations in the amount of inorganic matter com-ing from the watershed but rather due to changes in TOM fluxes.

Contrary to sediment cores of reference lakes, Lambda profilesin sediment cores of Matagami and Dickson lakes present markedvariations (Fig. 6) suggesting shifts in the ratio of allochthonous vs.autochthonous sedimentary organic matter. Matagami Lake’s sed-iment core presents sporadic and important variations of theLambda indicator suggesting that specific events induced tempo-rary variations in sedimentary OM fluxes trough the time. Becausethose changes are observed before and after the onset of the indus-trial era, we cannot attribute those modifications to either naturalevents or anthropogenic activities over the last 100 years. TheLambda profile of Dickson Lake shows a modification of the indica-tor occurring at 20 cm depth (Fig. 6) which, by extrapolation of thesedimentation rate, would correspond to the year 1600. This mod-ification is sudden and permanent suggesting an important salientevent, which modified the sedimentary regime of OM in the lake.Similar observations have been reported in other lakes (Hyodoet al., 2008; Louchouarn et al., 1993; Teisserenc et al., 2013). More-over, because this event happened before the European settlementof this region, we can assume that it corresponds to a major natural

event. In each perturbed lake, Lambda and Sigma 8 core profiles aresimilar suggesting that the variations observed for the Lambdaindicator are due to variations in TOM fluxes coming from the wa-tershed and not to fluctuations of the primary production into thewater column of the lake (Houel et al., 2006; Rezende et al., 2010).

4.2. Impact of mining and logging activities on TOM fluxes: couplinglignin biomarkers and a GIS approach

Using a GIS approach allowed us to confirm that the two refer-ences lakes have not been intensely exposed to logging and/ormining activities since 1979 whereas Waswanipi, Chibougamau,Matagami, Dickson and Des Jardins lakes watersheds have beenwidely logged and/or used by mining industries over the same per-iod of time (Moingt et al., 2013).

4.2.1. Impact of logging activitiesLignin biomarkers in lake sediments reflect the nature of vege-

tation cover and soil types in the lake watershed and thus can beused to identify the impact of logging activities on TOM fluxesreaching the lakes (Brandenberger et al., 2011; Gordon and Goñi,2003; Hedges and Mann, 1979; Louchouarn et al., 1999). Sedi-ments impacted by logging activities are characterized by higherLambda values than those of pristine lakes because of increasedtransfers of TOM from watersheds to aquatic systems in deforestedareas (Brandenberger et al., 2011; Loh et al., 2012). Fig. 7 presentsthe mean Lambda values measured in sediments covering the last100 year period for the studied lakes. Des Jardins E. and W. lakesare not considered because the difference in latitude between

Fig. 4. Total 210Pb activity vs. depth (A) and natural logarithm of unsupported 210Pb activity vs. depth (B) in Matagami Lake. 137Cs activity vs. depth (C) and 137Cs activity and210Pb age vs. depth (D) in the sediment core of Matagami Lake.

Fig. 5. Lambda (full black rhombs) and Sigma 8 (white diamonds) indicators vs. depth in reference lakes. Error bars are not visible at this scale. The straight black linesrepresent an age of 100 years for each core based on 210Pb dating.

40 M. Moingt et al. / Applied Geochemistry 41 (2014) 34–48

those lakes and reference lakes could explain a difference in the re-gime of the lakes as suggested by the high Sigma 8 values encoun-tered for Desjardins Lakes (Fig. 6). Sediments of Waswanipi andChibougamau lakes present mean Lambda values over the last100 year period similar to those measured in reference lakes(Fig. 7) and not different from the Lambda values measured in

the part of the core older than 100 years (Fig 6). Those observationssuggest minor impact of logging activities on TOM fluxes at thewatershed scale despite the fact they have been intensely impactedby logging activities since 1979 (Moingt et al., 2013). In contrast,sediments of Dickson and Matagami lakes display mean Lambdavalues over the last 100 years significantly higher than those of

Fig. 6. Lambda (full black rhombs) and Sigma 8 (white diamonds) indicators vs. depth in lakes presenting various types of anthropogenic perturbations occurring in theirwatersheds. Error bars are not visible at this scale. The straight black lines represent an age of 100 years for each core based on 210Pb dating.

M. Moingt et al. / Applied Geochemistry 41 (2014) 34–48 41

reference lakes suggesting at first that logging activities in theselakes may have increased TOM fluxes reaching the lake. However,in Dickson Lake, there is no difference between measured Lambdavalues in recent sediments and in sediments older than 100 yearsonce again suggesting a minor impact, at the watershed scale, oflogging activities on the amount of TOM reaching lake sediments.

In Matagami Lake we can observe strong variations in theLambda values both in recent sediments and in sediments olderthan 100 years. In the last 100 years, we can observe two majorpeaks of Lambda at ten and three centimeters. The beginning ofthe deeper peak corresponds to an event older than 100 yearsand then could not be related to logging activities which start in

Fig. 7. Mean Lambda values and associated standard deviation measured insediments covering the last 100 year period for the lakes of the study based on 210Pbdating. Gray area distinguishes reference lakes from the others lakes of the study.

Fig. 8. Mean S/V ratio values and associated standard deviation measured insediments covering the last 100 year period for the lakes of the study based on 210Pbdating. Gray area distinguishes reference lakes from the others lakes of the study.

Fig. 9. Mean P/(V + S) ratio vs. mean 3,5-Bd/V ratio values and associated standarddeviation measured in sediments covering the last 100 year period (full blackrhombs) and the period before the industrial era (white squares) for the lakes of thestudy based on 210Pb dating.

42 M. Moingt et al. / Applied Geochemistry 41 (2014) 34–48

the middle of the 1960s in the Matagami region. The beginning ofthe second peak corresponds to the nineties which can beassociated with the most intense period of forest cutting in theMatagami Lake watershed based on GIS results (Moingt et al.,2013). Then, due to the high natural variability of the Lambda val-ues, it is not possible to definitively relate any peak to anthropo-genic activities.

Logging activities are known to trigger the lixiviation of TOMnormally found in soils (Farella et al., 2001). Transfers of TOM fromlogged watersheds to aquatic systems are then characterized bylower S/V signatures, characteristic of humified TOM after pedo-genesis (Teisserenc, 2009). Sediments in lakes impacted by loggingactivities in their watersheds should then also record lower S/V ra-tios than those of pristine lakes. Fig. 8 presents the mean S/V ratiovalues measured in sediments covering the last 100 year period forthe lakes of the study. Those values are low when compared to S/V

ratios of pure angiosperm sources (Goñi and Hedges, 1992; Hedgesand Ertel, 1982; Hedges and Mann, 1979), suggesting a significantcontribution of gymnosperm species. However, Teisserenc (2009)showed that a comparison of S/V ratio in sediments to S/V ratioin soils would be a better tool for historical reconstruction of ter-restrial inputs to aquatic systems since soils are a better integratorof vegetation cover diversity, degradation and adsorption pro-cesses affecting TOM composition (Hernes et al., 2007; Louchouarnet al., 1999; Prahl et al., 1994; Sánchez-García et al., 2009). Ourmean S/V ratio values for the last 100 year period are in agreementwith the values reported by Teisserenc (2009) in boreal upper soilssuggesting then that the organic horizon of soils is the major con-tributor of sedimentary TOM. It is also suggesting that in borealecosystems, although the landscape is dominated by gymno-sperms, angiosperms are important contributors to sedimentaryTOM since gymnosperms do not produce syringyls (Hedges andMann, 1979; Hedges and Parker, 1976). Unfortunately, because ofthe strong variability of S/V values in the sediment core, this ligninindicator cannot be used to confirm that Matagami Lake sedimentsare actually significantly impacted by logging activities occurringin their watershed.

P/(V + S) and 3,5-Bd/V ratios are used to estimate the level ofdegradation and the degree of maturation of TOM respectively(Dittmar and Lara, 2001; Houel et al., 2006). Moreover, 3,5-Bd/V ra-tios could be used to distinguish TOM from organic and inorganicsoil horizons (Houel et al., 2006). Fig. 9 presents mean P/(V + S)and 3,5-Bd/V ratios measured in sediments over the last 100 yearperiod for the lakes of the study. Chibougamau Lake probably doesnot fit into the regression observed in Fig. 9 for two reasons: (i)p-hydroxyphenols amount measured in sediments over the last100 years is similar to the one measured in the other sedimentcores, and (ii) both syringyl and vanillyl contents are the lowestcompared to the other lakes cores examined in this study(0.06 ± 0.04 and 0.24 ± 0.04 mg/100 mg of OC respectively). Chibo-ugamau Lake has the lowest sedimentation rate of all the lakes ofthis study which could explain really low levels of syringyl phenolssince they are more sensitive to degradation processes than theothers phenol families (Houel et al., 2006; Opsahl and Benner,1995). Both Matagami and Dickson lakes present significantly low-er P/(V + S) and 3,5-Bd/V ratios suggesting that sedimentary TOMin those lakes comes from the upper soil organic horizon (Houelet al., 2006; Teisserenc, 2009), which is in agreement with the fact

M. Moingt et al. / Applied Geochemistry 41 (2014) 34–48 43

that draining processes are more superficial in clear-cut land incomparison to natural forest cover (Farella et al., 2001). However,despite this observation, in the case of large lakes boreal ecosys-tems, at a watershed scale our data suggest that logging activitiesglobally does not seem to have a noticeable impact on sedimentaryTOM signatures and fluxes. Then large-scale watersheds showsome ‘‘buffering’’ capacity to logging disturbances since about30% of cut in a watershed is needed to start showing an impacton sedimentary TOM (Moingt et al., 2013). Our study is based onlarge lakes with large watersheds (watershed areas comprised be-tween 58 and 1084 km2), then the system’s inertia is most likelyresponsible for this ‘‘buffering’’ capacity since it increases theamount of time needed for the perturbation to reach the system.Indeed, during the transfer of organic matter from the terrestrialsystem to the aquatic system, several processes could contributeto mask and/or dilute the signal of the perturbation (e.g. OM deg-radation, OM incorporation into soils, fractionation of the TOMpool).

4.2.2. Impact of mining activitiesThe presence of mine tailing in the watershed normally in-

creases the amount of inorganic component of sediment reachingthe lake (Petit et al., 2011). This increase should then induce a de-crease in the Sigma 8 in lake sediments by diluting the TOM signalin the sediment sample. However, Fig. 6 shows that for the fourlakes presenting mining activities in their watersheds, Sigma 8 isstable or slightly increases. Moreover, soil erosion normally trig-gers higher rates of transfer of dissolved and particulate nutrients

Fig. 10. THg (ng/g) and Sigma 8 (mg/10 g of sample) vs. depth in Matagami and Dickson210Pb dating.

Fig. A1. Total mercury contents vs. depth in reference lakes (i.e. Ouescapis and Rodayer210Pb dating. Figure modified from Moingt et al. (2013).

from the watershed to the water column (Fraser et al., 1999; Klee-berg et al., 2008; Sharpley et al., 2001). These additional nutrientloads then enhance the production of aquatic biota. A change inthe primary production of the lake can be detected by the studyof the Lambda indicator since it is proportional to the allochtho-nous/autochthonous ratio of OM in a sediment sample. Over thelast 100 year period, only Matagami Lake presents variations inthe Lambda indicator (Fig. 6). In this case, Lambda variations alongthe core cannot be associated with changes in primary productionas explained previously in Section 4.1. The onset of mining activi-ties in the watershed could also induce a change in TOM sources.First, excavation processes bring to the surface important amountsof TOM from deeper soil horizon with lignin signatures impactedby pedogenesis (i.e. high P/(V + S) and 3,5-Bd/V ratios). Second,mining activities enhance drainage and erosion rates of surface soilhorizons (Mäkinen et al., 2010) characterized by less degradedTOM (i.e. low P/(V + S) and 3,5-Bd/V ratios). However, studied lakeswith mining activities in their watersheds do not present anychange of the P/(V + S) vs. 3,5-Bd/V signatures from the pre-indus-trial period to present (Fig. 9). Biomarkers data then suggest that inlarge lakes with large watersheds, the impact of mining activitieson TOM fluxes is negligible when considering sediment cores inte-grating the whole watershed.

4.3. Combined transfers of TOM and Hg from watersheds to lakes

210Pb dating allows us to confirm that for all the lakes of thepresent study the significant increase in THg contents in sediment

lakes sediment cores. The straight black line represents an age of 100 years based on

lakes). The straight black lines represent an age of 100 years for each core based on

Fig. A2. Total mercury contents vs. depth in lakes presenting various type of anthropogenic perturbations in their watershed. The straight black lines represent an age of100 years for each core based on 210Pb dating. Figure modified from Moingt et al. (2013).

44 M. Moingt et al. / Applied Geochemistry 41 (2014) 34–48

coincide with the beginning of the industrial era (Appendix A,Figs. A1 and A2).

THg contents increase in the last 100 years is not necessarilyassociated with an increase in the Lambda and Sigma 8 indicators(Figs. 5 and 6), which would indicate an increase in TOM inputsreaching lake sediments. Then, the results of the present studyshow that in large boreal lake ecosystems there is no clear evi-dence of an impact of land-uses occurring in the watershed on bothsedimentary TOM and THg contents. This lack of correlation be-tween the evolution of Lambda and Sigma 8 indicators and THgcontents along a sediment core could be due to the fact that theamount of TOM naturally transferred to the lake is widely suffi-cient to scavenge and bind the amount of Hg being deposited fromthe atmosphere even with this amount increasing over the last150 years due to the industrial era while TOM fluxes remain rela-tively constant. However, this observation is not valid for Mata-gami and Dickson lakes (Fig. 10).

Considering Matagami Lake, sporadic variations in the Sigma 8indicator are observed before and after the onset of the industrialera. Before that era, Sigma 8 variations are not correlated to THgfluctuations whereas over the last century, trends in Sigma 8 andTHg contents do correlate (r = 0.93, p = 0.0187). This pattern sug-gests that THg peaks measured in sediments over the last 100 yearperiod are due to transfers of recently deposited atmospheric Hgfrom the watersheds to the lakes. Moreover, because an increasein TOM fluxes induces an increase in THg contents in sediments,any contemporary watershed perturbation (natural and/or anthro-pogenic) influencing TOM fluxes will affect sediment THg levels.

In Dickson Lake, an important increase in the Sigma 8 indicatorat 20 cm depth is observed and corresponds to a sharp increase inTHg in sediments (r = 0.87, p < 0.0001). An extrapolation of 210Pbdating results allows us to situate this event �400 years ago andthus conclude that this anomaly corresponds to a natural event.Moreover, after this event, both Sigma 8 and THg did not comeback to their initial levels suggesting a drastic change in the wa-tershed lake regime after natural flooding or landslide for example.Teisserenc et al. (2013) observed the same pattern in flooded lakesfor both Lambda and THg in sediment cores and concluded thatthis pattern was also evidence of a close association betweenTOM and THg loading. Dickson Lake is also a good example show-ing the importance of sedimentation rates to better understand theHg cycle in boreal lakes ecosystems. In this lake, the ratio betweenTHg baseline concentration and THg peak at the surface is 15 (i.e.ASEF = 15). However, knowing that most of the THg increase ob-served in the sediment core is pre-industrial era, probably due toa natural change of the regime of the lake, for a representative ASEFwe have to change the THg baseline content previously considered.The new ASEF becomes 3.2, which is more typical of other NorthAmerican lakes reported in the literature (Kamman and Engstrom,2002; Lucotte et al., 1995; Swain et al., 1992).

5. Conclusion

Over the last decade, several publications based on the study ofsmall lakes point to anthropogenic perturbations occurring in the

Table A2Mercury characteristics in lake sediment cores. Data are from Moingt et al. (2013).

Lake THg baseline insediment (ng/g)

THg peak insediment(ng/g)

ASEF EHg(ng/g)

Des Jardins 169 364 2.2 196Ouescapis 33 89 2.7 56Rodayer 24 67 2.8 43Waswanipi 40 98 2.5 58Chibougamau 40 150 3.8 110Dickson 6 90 15 84Matagami 16 87 5.4 (surface) 71

15.0 (peak)

Table A1Biogeomorphological characteristics of the lakes of the present study as presented in Moingt et al. (2013). %TWA = percentage of the total watershed area, Gr = gravel roads,L = logging activities, U = urbanization, G = golf.

Lake Localizationlat./long.

Altitude(m)

Lake area(km2)

Drainagearea (km2)

DA/LA Meanslope (%)

Year Water(%TWA)

Coniferous(%TWA)

Mixed forest,deciduous >50%(TWA%)

Unforested(%TWA)

Number ofmines in thewatershed

Majordisturbances

Des Jardins 46�390N78�150W

320 15.02 207.39 13.81 5.85 2001 9.62 0.00 33.66 3.57 x Gr, L2004 10.04 0.00 18.46 0.372010 9.56 0.00 43.36 2.27

Ouescapis 50�150N76�600W

278 36.41 107.94 2.96 4.84 1979 0.00 0.00 7.01 2.35 x Gr1990 0.00 0.00 1.80 1.802010 0.02 0.00 5.12 13.98

Rodayer 50�510N77�410W

249 22.31 57.7 2.59 1.34 1979 1.93 0.00 13.94 3.65 x Gr1990 2.12 0.00 5.33 0.952010 2.43 0.00 3.27 2.44

Waswanipi 49�330N76�270W

267 200.69 1483.66 7.39 2.02 1979 1.41 0.00 12.05 7.35 3 Gr, L1990 0.65 0.00 5.47 20.382010 1.36 0.00 7.16 15.61

Chibougamau 49�500N74�130W

378 201.06 769.87 3.83 4.45 1999 8.68 0.00 18.66 16.31 14 Gr, L, U2002 8.76 0.00 21.13 5.822009 8.61 0.00 19.17 10.18

Dickson 49�380N75�110W

342 10.00 227.59 22.76 3.23 1979 0.85 0.00 11.63 5.65 1 Gr, L1990 1.58 0.00 2.63 12.652010 1.30 0.00 10.37 7.78

Matagami 49�500N77�370W

248 209.86 951.36 4.53 2.84 1979 2.15 0.00 12.94 8.72 7 Gr, L, U, Go1990 0.24 0.00 8.46 27.272010 1.14 0.00 12.15 8.28

M. Moingt et al. / Applied Geochemistry 41 (2014) 34–48 45

watershed (e.g. logging and mining activities) as a contributor tothe increasing amount of THg reaching lakes sediments (Garciaand Carignan, 2000, 2005; Porvari and Verta, 2003; Porvari et al.,2003). In contrast to these findings on watersheds for small lakes,the present study demonstrates that the impact of major anthropo-genic activities (e.g. extensive logging, mining) occurring in thewatershed of large boreal lakes is not necessarily discernible inHg and TOM concentrations in sediment cores integrating thewhole watershed. Indeed, in most of the lakes with anthropogenicperturbations in their watersheds, THg profiles are typical of atmo-spheric/deposition impacted sediments (i.e. ASEF around 3)whereas the sedimentary lignin signature does not change afterthe beginning of those activities in the watershed. However, inhighly impacted watersheds (e.g. forest clearing in the order of30% the watershed surface over 10 years) both Hg and organic mat-ter contents can be influenced. This study indicates that large-scalewatersheds are characterized by some ‘‘buffering’’ capacity to land-use disturbance, which would be more likely due to the system’sinertia since it increases the amount of time needed for a perturba-tion to reach the system. Indeed in large watershed boreal ecosys-tems, THg and TOM are more likely to be impacted by processes,such as OM degradation, OM incorporation into soils, fractionation

of the TOM pool, incorporation of Hg into wetlands, Hg adsorptionon the soil mineral phase, which could contribute to mask and/ordilute the signal of the perturbation.

Irrespective of the presence or absence of an impact from land-use, this study underlines the importance of TOM, based on ligninbiomarkers results, as a vector of Hg transfer from the watershedto the receiving lakes since sedimentary Hg enrichment seems tobe directly proportional to the amount of TOM coming from thewatershed (Sigma 8 indicator). This is critical since TOM flows nat-urally from watershed soils and has thus contributed to a markedincreased accumulation of atmospheric Hg in lake sediments. Then,any natural and/or anthropogenic modification increasing TOMfluxes from the watershed to the lake can increase the amount ofHg reaching the lake, this supplementary Hg having the potentialto be methylated and then incorporated into the aquatic foodchain. Because most of the atmospheric/deposited Hg is associatedwith ground vegetation and with soils (Hintelmann et al., 2002;Ouellet et al., 2009) a decrease in Hg emissions should ultimatelylead to a decrease in the amount of Hg transferred from the wa-tershed to the lake. A solid knowledge of both bio-morphologicalparameters (e.g. mean slope, vegetation cover; Moingt et al.,2013) and TOM inputs to the lake could help to identify water-sheds at higher risk of exporting Hg to the lake in order to limitthe human health risk by fish consumption.

Acknowledgments

This project was financed through the CARA (Clean Air Regula-tory Agenda) program of Environment Canada. We would like tothank Sophie Chen for her assistance in the laboratory and Jean-Sébastien Beaulne for his help during the sampling work.

Appendix A. Appendix

See Figs. A1 and A2 and Tables A1–A3.

Table A3Primary lignin–phenol results. In each core, the gray area corresponds to the portion of the core younger than 100 years. allo = allochthonous, auto = autochthonous.

46M

.Moingt

etal./A

ppliedG

eochemistry

41(2014)

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M. Moingt et al. / Applied Geochemistry 41 (2014) 34–48 47

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