PRIMARY RESEARCH PAPER

Paleolimnological assessment of limnological changein 10 lakes from northwest Saskatchewan downwindof the Athabasca oils sands based on analysis of siliceousalgae and trace metals in sediment cores

Kathleen R. Laird • Biplob Das • Melanie Kingsbury •

Melissa T. Moos • Sergi Pla-Rabes • Jason M. E. Ahad •

Brendan Wiltse • Brian F. Cumming

Received: 15 January 2013 / Revised: 12 June 2013 / Accepted: 20 July 2013 / Published online: 4 August 2013

� Springer Science+Business Media Dordrecht 2013

Abstract The extraction of bitumen from the Ath-

abasca oil sands is rapidly expanding, and emission of

sulphur and nitrogen oxides has substantially increased.

To determine whether lakes downwind of this devel-

opment in northwest Saskatchewan have been detri-

mentally impacted since development of the oil sands, a

paleolimnological assessment of ten lakes was carried

out. Analysis of diatom valves and inferences of

diatom-inferred pH indicated that emissions have not

resulted in widespread chronic acidification of acid-

sensitive lakes *80–250 km east and northeast of the

oil sands development around Fort McMurray and Fort

Mackay. However, one of the closest sites to the

development indicated a slight decline in diatom-

inferred pH, but the two next closest sites, both of which

had higher alkalinity, did not show any evidence of

acidification. There were also no consistent trends in the

concentration or flux of total or individual priority

pollutants including lead, mercury, copper, zinc and

vanadium. The sedimentation rates in most lakes

increased since the mid-1900s, along with increased

flux of both diatoms and scaled chrysophytes. Subtle

changes in the species assemblages of diatoms and

increased flux of diatoms and chrysophyte scales are

consistent with recent climate change in this region.

Keywords Lake acidification � Climate

change � Diatoms � Chrysophyte scales � Trace

metals � Athabasca oil sands � Saskatchewan

Handling editor: Jasmine Saros

Electronic supplementary material The online version ofthis article (doi:10.1007/s10750-013-1623-5) contains supple-mentary material, which is available to authorized users.

K. R. Laird (&) � M. Kingsbury � B. Wiltse �B. F. Cumming

Paleoecological Environmental Assessment and Research

Laboratory (PEARL), Department of Biology, Queen’s

University, Biosciences Complex, Kingston,

ON K7L 3N6, Canada

e-mail: lairdk@queensu.ca

B. Das

Saskatchewan Ministry of Environment,

3211 Albert Street, Regina, SK S4S 5W6, Canada

M. T. Moos

Lake Simcoe Region Conservation Authority,

120 Bayview Pkwy., Newmarket, ON L3Y 4X1, Canada

S. Pla-Rabes

CEAB-CSIC, Ctr Adv Studies Blanes, LimnolGrp,

17300 Blanes, Girona, Spain

S. Pla-Rabes

Centre for Ecological Research and Forestry Applications

(CREAF), 08193 Cerdanyola del Valles, Spain

J. M. E. Ahad

Geological Survey of Canada, Natural Resources Canada,

490 rue de la Couronne, Quebec, QC G1K 9A9, Canada

123

Hydrobiologia (2013) 720:55–73

DOI 10.1007/s10750-013-1623-5

Introduction

The extraction of bitumen to produce usable oil from

Alberta’s oil sands began in 1967; but due to the high

cost of production and lack of appropriate technology,

the development did not start in earnest until the 1990s

(Schindler, 2010). Today, production is increasing

nearly exponentially (Schindler, 2010). Recent

increases in the price of conventional oil and improved

extraction techniques have led to this rapid expansion

(Parsons et al., 2010a). There are numerous potential

environmental impacts to the surrounding land, water

and air with this development. As the industrial

footprint in the Athabasca oil sands continues to

expand, emissions of carbon dioxide, sulphur oxides,

nitrogen oxides and other pollutants will continue to

increase (Aherne & Shaw, 2010; Kelly et al., 2010;

Parsons et al., 2010b; Schindler, 2010).

Recent research analysing snow-pack samples

indicates that there have been significant releases of

pollutants to the Athabasca River and its watershed as

the result of emissions from the Athabasca oil sands

development (Kelly et al., 2009, 2010). Concentra-

tions of the US Environmental Protection Agency’s

priority pollutant elements (PPEs) under their Clean

Water Act were found to be higher in more disturbed

watersheds versus those further away from develop-

ment. Concentrations of seven of the thirteen PPE—

cadmium, copper, lead, mercury, nickel, silver and

zinc—were found to have exceeded Canada’s guide-

lines for aquatic life (Kelly et al., 2010). Further

indication of the industrial footprint was found in

increased concentration of polycyclic aromatic hydro-

carbons (PAHs) in the snow-pack samples (Kelly

et al., 2009), as well as increased concentrations in

lake sediment cores adjacent to the main Athabasca oil

sands mining operations coincident with increased

development (Kurek et al., 2013). With the expansion

of the oil sands, the process of upgrading bitumen to

crude oil will be a continuing source of PPEs and

PAHs to the environment.

Biological assessment of potential impacts on lakes

from the increasing emissions of sulphur and nitrogen

compounds from the expansion of the oil sands has

begun in regions closest to the development of the oil

sands, as these areas are expected to be the most

susceptible to adverse impacts (Hazenwinkel et al.,

2008; Curtis et al., 2010; Parsons et al., 2010a). In

eastern North America, emission of sulphur and

nitrogen compounds has been clearly linked to acid

deposition and lake acidification (e.g. Charles & Smol,

1990; Sullivan et al., 1990; Cumming et al., 1994).

Deposition of nitrogen from fossil emissions could

also lead to nutrient enrichment in aquatic ecosystems

resulting in increased algal production (Bergstrom &

Jansson, 2006; Baron et al., 2011; Greaver et al.,

2012). Paleolimnological assessment of lake sedi-

ments provides a means of demonstrating this link by

establishing baseline conditions and range of natural

variability that is typically not available from other

monitoring records. The study of remains of diatoms is

a well-established technique for inferring change in

lake-water pH through time based on the diatom

assemblages preserved in dated lake sediment cores

(Flower & Battarbee, 1983; Battarbee et al., 1985;

Smol et al., 1986).

Studies of potential impacts on lakes more down-

wind of emissions from the Athabasca oil sands have

not been extensively assessed. A survey of the water

chemistry from 259 headwater lakes within 300 km of

Fort McMurray concluded that acid sensitivity based

on lake-water acid neutralising capacity (ANC) varied

by ecoregion with lakes in the region located adjacent

to oil sands being the least sensitive (the Mid-Boreal

Upland ecoregion), and the Churchill River upland

and Athabasca plains being more sensitive (Scott

et al., 2010). Although water chemistry characteristics

and process-based models (Whitfield et al., 2010)

suggest that anthropogenic acidification have likely

not occurred or are in their early stage, a longer

temporal perspective based on empirical changes in

indicators of acidification is prudent. In this study, we

use a paleolimnological approach to evaluate if recent

development and expansion of the Athabasca oil sands

in Alberta has resulted in lake acidification or change

in metal fluxes of pollutants to the lakes through

atmospheric deposition in northern Saskatchewan. We

examine this through the analysis of sediment records

from ten lakes in Saskatchewan located to the east and

northeast of the main area of the Athabasca oil sands

mining operations. The objectives of this study are to

establish baseline conditions prior to 1980 when

development of the oil sands began to expand. The

biological baseline is accomplished through the anal-

ysis of diatom assemblages and inferred pH changes.

Fluxes of elements, known to be linked to oil

56 Hydrobiologia (2013) 720:55–73

123

production and bitumen upgraders released into the

atmosphere (Kelly et al., 2010) were estimated

from the sediments to determine whether emissions

of these pollutants have increased through time in this

region.

Materials and methods

Study sites

Ten lakes were selected in northern Saskatchewan

(Fig. 1), northeast of oil sands development near Fort

McMurray, across the Boreal plain (Mid-Boreal

Upland ecoregion, lakes 8D, 10W and 10G) and the

Boreal Shield (lake 13C, in the Churchill River Upland

ecoregion, and the remainder on the Athabasca plain).

A list of candidate lakes were selected based on a water

chemistry survey of 259 headwater lakes sampled

during 2007–2008 within*300 km of Fort McMurray

(Scott et al., 2010). Criteria for selection of the study

lakes included: being relatively small undisturbed

headwater lakes in Saskatchewan (depths between 4

and 15 m, with lake surface areas \200 ha); with

relatively low dissolved organic carbon (DOC) values

(\6 mg l-1) and relatively low alkalinity

(ANC \ 200 leq l-1). By using these criteria we

hoped to identify lakes that were very-sensitive

(ANC \ 50 leq l-1) or sensitive (ANC \ 200

leq l-1) to acidic deposition (Scott et al., 2010) that

were not complicated by other forms of anthropogenic

(roads, mining, forestry) or natural (e.g. fire) distur-

bance in their catchments, but located at varying

distances from the predominant anthropogenic emis-

sions associated with the mining operations of the

Athabasca oil sands. The majority of the lakes sampled

(n = 6) were located on the Athabasca plain, a region

that was identified as the most acid-sensitive ecoregion

(Scott et al., 2010). The study lakes can be character-

ized as small (80% with surface areas \50 ha, and

depths between 4 and 7 m), slightly acidic to circum-

neutral lakes (pH range from 6.3 to 7.3), low conduc-

tivity (90%\20 lS cm-1), with low concentrations of

DOC (90% of DOC values range from 2.5 to

5.9 mg l-1; exception of 8D at 11.6 mg l-1), that are

relatively low in nutrients (total phosphorus 5–18

lg l-1) (Table 1). Four lakes can be categorized as very

sensitive (ANC values \50 leq l-1; lakes 9D, 8H, 7A,

12D), four lakes as sensitive (ANC

values \ 100 leq l-1; lakes 15F, 10W, 13C and 15J),

and two lakes as more buffered (lakes 8D and 10G).

Fieldwork was based out of Axe Lake Camp (Oil

sands Quest Inc.) and lakes were accessed by

helicopter. Given the remoteness of the study lakes,

bathymetric information and watershed characteristics

are limited.

Sediment coring and sampling

Sediment cores were retrieved from all lakes with a

Glew gravity corer (internal diameter *7.62 cm,

Glew et al., 2001) between March 4th and 6th, 2010.

The ice surface was used as a platform and an auger

was used to drill through the ice. The length of cores

retrieved ranged from 24 to 60 cm (average *40 cm)

and were sectioned into 0.25-cm intervals (Glew et al.,

2001) upon return to the base camp.

Core samples were shipped to Queen’s University,

where they were stored in a 4�C cold room. The weight

of all sediment samples was determined to calculate

the total weight of sediment prior to sub-sampling for210Pb, diatoms, chrysophytes, mercury and metals.

Twenty intervals were sub-sampled for diatoms,

chrysophytes, metals and mercury every 1 cm from

0.25 to 20 cm.

Radiometric dating

Twenty intervals were prepared for 210Pb analysis and

counted using gamma spectroscopy at PEARL Queen’s

University, with samples more closely spaced in the

higher activity region (every 1 cm for the top 10 cm)

and at coarser intervals down core (every 2 cm from

12–28 cm). Samples were dried in a freeze drier and the

dry weight and percent moisture in each sample was

determined. Dried sediment was weighed into a plastic

tube for gamma spectroscopy. The samples were then

sealed with epoxy and allowed to sit for 2 weeks in

order for 214Bi to equalize with 226Ra, the parent isotope

of 210Pb. Activities of 210Pb, 137Cs, and supported 210Pb

(via 214Bi) were determined for each sample using

gamma spectroscopy following the procedures outlined

in Schelske et al. (1994). Unsupported 210Pb activities

were used to estimate the chronology of the cores using

the constant rate of supply (CRS) model (Appleby &

Oldfield, 1978), using the computer programme devel-

oped by Binford (1990), or run in R ver. 2.8.1. (Jeziorski

& Thienpont, PEARL, unpublished programme).

Hydrobiologia (2013) 720:55–73 57

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Fig. 1 Location of the 10 study lakes in northwest Saskatchewan

Table 1 Summary of physical and chemical lake characteristics

Lake Surface

area (ha)

Coring

depth (m)

pH Alk

(mg l-1)

Specific

conductance

(lS cm-1)

DOC

(mg l-1)

Chl

a (lg l)

TP

(lg l-1)

TN

(lg l-1)

9D 3 9 6.4 1.5 6.7 2.5 1.1 5.0 318

8H 5 4 6.3 2.1 8.2 5.6 4.6 14.5 576

7A 20 4 6.3 1.6 7.1 5.6 2.6 10.0 501

8D 22 4.3 6.9 11.1 25.3 11.6 8.2 18.0 942

15F 36 4.4 6.7 4.3 17.5 4.8 2.8 10.5 618

10G 38 5.5 7.3 17.9 34.6 3.2 4.3 13.0 474

10W 39 6.8 6.5 3.8 11.7 5.9 3.2 9.0 674

12D 43 6.4 6.5 2.7 12.8 2.5 9.6 12.5 295

13C 146 4.8 6.7 4.5 14.0 5.8 7.0 11.5 387

15J 243 6.3 6.9 5.0 13.9 3.8 2.6 8.0 313

Lakes are ordered according to increasing surface area

Alk alkalinity, DOC dissolved organic carbon, Chl a chlorophyll a, TP total phosphorus, TN total nitrogen

58 Hydrobiologia (2013) 720:55–73

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Unsupported 210Pb was calculated by subtracting

supported 210Pb, except for two lakes (8D, 10G) where

total 210Pb did not reach supported level by 28 cm. In

these cases, unsupported 210Pb was calculated by

subtracting an estimate of background activity of

3 dpm g-1 from the total 210Pb activity.

Diatom and scaled chrysophyte analyses

For each core, *0.2–0.3 g of wet sediment was sub-

sampled and placed in 20-ml glass vials to which a 1:1

mixture by molar weight of concentrated nitric (HNO3)

and sulphuric (H2SO4) acid was used to remove organic

matter. The samples were allowed to settle for 24 h

before the acid above the sample was removed, and the

sample was rinsed with distilled water. This procedure

was repeated until the sample had the same pH as the

distilled water (approximately eight rinses). Four

successive dilutions for each sample were pipetted

onto coverslips ensuring that each sample was well

mixed. Samples on the coverslips were air-dried

overnight, then heated on a warming plate to remove

any remaining moisture, and subsequently mounted

with Naphrax� onto glass microscope slides. Diatoms

were identified and counted along transects on the

prepared slide using a Leica (DMRB model) micro-

scope fitted with a 1009 fluotar objective (numerical

aperture of objective = 1.3) and using differential

interference contrast optics at 1,0009 magnification.

Approximately 400 diatom valves were enumerated per

slide. Diatoms were identified down to the species level

or lower, using the following taxonomic references:

Krammer & Lange-Bertalot (1986, 1988, 1991a, b),

Lange-Bertalot & Melzeltin (1996), Camburn &

Charles (2000) and Fallu et al. (2000).

Concentration of diatoms was determined for all

samples using methods outlined in Battarbee & Kneen

(1982). An aliquot of a known concentration of

microspheres was added to each of the diatom

samples, prior to settling on coverslips. The micro-

spheres were enumerated along with the diatoms and

used to calculate estimates of number of diatoms per

gram dry weight of sediment.

Chrysophyte scales were identified and enumerated

using the slides prepared for diatom analysis. Due to

the low concentration of scales, only 100–200 were

counted per sample where possible (typically with

microsphere counts from *1,000–3,000, and up to

5,000). In many samples the concentration of scales

was too low to count. Standard taxonomic references

were used and are outlined in Cumming et al. (1992).

Concentrations were calculated using microspheres in

relation to chrysophytes, using the procedures

described above. The flux rates of diatoms and

chrysophyte scales were calculated by multiplying

the concentration (# per g dry weight) by the

sedimentation rate (g cm2 year-1) obtained from the

CRS chronology model.

Trace metal analysis

For the metal analysis, 0.5 g each of freeze-dried

sample was digested using 2% HNO3 and the metal

contents were determined at the Environmental Anal-

ysis Laboratory, Saskatchewan Research Council

(Saskatchewan Canada) using the ICP-MS scan. Total

metal concentrations of Ag, Al, As, B, Ba, Be, Cd, Co,

Cr, Cu, Fe, Mn, Mo, Ni, Pb, Sb, Se, Sn, Sr, Ti, U, V and

Zn were reported as lg g-1 of dry sediment. Flux rates

were calculated by multiplying the concentration

(lg g-1) by the sedimentation rate (g cm2 year-1)

obtained from the CRS chronology model.

Mercury analysis was carried out at the Analytical

Services Unit at Queen’s University on approximately

100 mg of freeze dried samples on a Dedicated

Mercury Analyser (DMA-80), using a CALA accred-

ited method [Analytical Services Unit, Method Num-

ber 12 (ASU 12-Mercury by DMA-80)]. In brief, the

solid samples are dried and thermally decomposed

with the aid of a hot catalyst bed in a continuous flow

of oxygen. The Hg vapours were trapped on a gold

amalgamator tube, and then were measured by vapour

atomic absorption spectrophotometry at 254 nm

(USEPA method 7473-mercury in solids and solutions

by thermal decomposition, amalgamation, and atomic

absorption spectrophotometry). The Hg detection

limits for this technique is 5 ppb.

Numerical analyses

Inferences of pH from the diatom assemblages in the

cores are based on a pH model developed from 47

freshwater lakes from across northern Alberta (Pla &

Curtis, 2006), and is the same model used in Hazen-

winkel et al. (2008) and Curtis et al. (2010). The

Hydrobiologia (2013) 720:55–73 59

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60 Hydrobiologia (2013) 720:55–73

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assemblage data used were all taxa greater than 1%. A

two component weighted-averaging partial-least-

squares model based on square-root transformed

relative percentage data was used. The bootstrapped

coefficient of determination (r2) of this model is 0.81,

with a RMSEP of 0.42 pH units. The performance of

this model was assessed through comparison of the

diatom-inferred pH of surface samples to the measured

pH of the study lakes, as well as comparisons to other

pH inference models (Hazenwinkel et al., 2008).

Hazenwinkel et al. (2008) indicated that the best

predictive powers of this model were between *pH 6

and 7.5, which incorporates all of our study lakes.

The main direction of variation in the diatom

assemblages were determined from a principal com-

ponents analysis (PCA) using non-transformed spe-

cies abundance data ([1%) and centred data by

variables using C2 v 1.6.3 (Juggins, 2003). PCA was

chosen to represent the main directions of variation of

the diatom assemblages because the gradient length in

an initial detrended correspondence analysis on each

of the study lakes was \1.5 standard deviation units.

Correlations between the estimated pH and the first

PCA axis scores were determined.

The diatom assemblage zones in the down-core

analyses were defined by a constrained cluster analysis

(Grimm, 1987) and verified with the broken-stick

method using optimal zonation to establish the order

of the splits and their significance using the programme

PSIMPOLL v 4.10 (Bennett, 1996). Those with the

highest significance are considered the ‘first order’ or

major zones, whereas those with lower significance are

sub-zones. Only the first-order zones are shown in Fig. 2

and are used as a framework for discussing the results.

Results

Chronology

The majority of lakes showed an exponential decay of

total 210Pb activity with depth (Fig. Supplemental 1,

S1), enabling the development of a strong depth-time

relationship. Background (or supported) levels of210Pb (where total 210Pb activity and total 214Bi

intersect) are reached by 12–16 cm in four lakes

(7A, 8H, 9D, 10W), in four other lakes (12D, 13C,

15F, 15J) between 20 and 26 cm and at greater than

the analysed portion in two lakes (8D, 10G) (Fig. S1).

The relationships between log unsupported 210Pb and

cumulative dry mass are strong (r2 = 0.85–0.95) and

highly significant in the sediment cores from six of the

10 study lakes. Three lakes (8D, 10G, 15J) have a

weaker relationship between log unsupported 210Pb

and cumulative dry mass (r2 = 0.54–0.68), but suffi-

cient chronological control for determining if recent

development of the Athabasca oil sands has had an

impact on the study lakes. Only lake 15F has a weak

relationship between log unsupported 210Pb and

cumulative dry mass (r2 = 0.24), and thus estimates

of the timing of changes in this record need to be

interpreted with caution.

Uncertainty of 210Pb estimated dates increases pre-

1900, thus only post-1900 analyses are shown for both

siliceous microfossils and metals. The number of

samples presented varies between the cores because

the range of time represented by the top 20 cm for

which siliceous microfossils and metals were analysed

varied due to the differences in the sedimentation rate

between lakes.

Diatom species assemblages

Approximately 50–100 taxa were identified within

each of the 10 study lakes; however, the majority of

these were rare, with only 5–14 taxa reaching abun-

dances greater than 5%. The majority of the assem-

blages for all lakes consisted primarily of benthic taxa,

with the exception of lake 8D where the planktonic

portion consisted of 30–66% (average of 48%) of the

assemblage. Two other lakes also had a fairly large

proportion of planktonic taxa, average of 33% (range

24–47%) in lake 7A and 25% (range 13–32%) in lake

15J. The dominant planktonic taxon in these three

lakes are Discostella stelligera (Cleve & Grunow)

Houk & Klee, with D. pseudostelligera (Hustedt)

Houk & Klee also present (Fig. 2). Other less abun-

dant planktonic or tychoplanktonic taxa include spe-

cies in the genus Aulacoseira (ambigua Simonsen,

lacustris (Grunow) Krammer, perglabra (Østrup)

Haworth, subarctica (Muller) Haworth).

Fig. 2 Percent abundance of the dominant diatom taxa of each

of the study lakes. Only taxa which are either dominant

([5–10%) or indicate a change are shown. Lakes are ordered

according to increasing change in species assemblages based on

PCA length (Table 2). Horizontal lines indicate the two zones

based on constrained cluster analysis (Grimm, 1987) and

optimal zonation (Bennett, 1996)

b

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A diversity of benthic taxa were common across the

lakes, with Staurosirella pinnata (Ehrenberg) Wil-

liams & Round (in the family Fragilaraceae Greville)

being one of the most common, reaching abundances

of 40–50% in two lakes (12D, 13C) and nearly 90% in

lake 10G (Fig. 2). Staurosira construens Ehrenberg

and S. construens v. venter (Ehrenberg) Hamilton

were also common Fragilaraceae, reaching abun-

dances of*20–40% in a few lakes (e.g. 8H, 9D, 12D).

These small Fragilaraceae can form large chains and

thus sometimes are found within the plankton, partic-

ularly in more shallow systems or in near-shore

environments.

Other common benthic taxa included Achnanthidi-

um minutissimum (Kutzing) Czarnecki, Brachysira

garrensis (Lange-Bertalot & Krammer) Lange-Berta-

lot and Navicula leptostriata Jørgensen which reached

abundances [20% in at least one lake. As with the

more abundant taxa, those taxa that reached abun-

dances of between 5 and 20% varied tremendously

between lakes, but with some having a more common

occurrence across the study lakes, such as Brachysira

brebissonii Ross, Chamaepinnularia mediocris (Kra-

sske) Lange-Bertalot, Naviculadicta subtilissima

(Cleve) Lange-Bertalot, Nitzschia perminuta (Gru-

now) Peragallo and Pinnularia interrupta W. Smith.

Scaled-chrysophyte species assemblages

Eight lakes had sufficient concentrations of scaled

chrysophytes to enumerate from 5 to 20 samples in

each core, with only lake 8D having sufficient

numbers for all 20 samples (Fig. Supplemental 2,

S2). Only lake 10G had no samples that had sufficient

concentrations, and lake 12D only had the very top

sample with sufficient numbers to enumerate (not

shown). The dominant taxon across all lakes was

Mallomonas crassisquama (Asmund) Fott, a common

taxa in circumneutral to alkaline lakes (Cumming

et al., 1992; Siver, 1995). In lakes 13C and 15F,

Synura sphagnicola (Korshikov) Korshikov was also

dominant, which is a colonial chrysophyte with a

lower pH optima at 6.4 (versus 7.1 for M. crassisqu-

ama) (Cumming et al., 1992). Other Mallomonas

present in many of the lakes were M. duerrschmidtiae

Siver, Hamer & Kling and M. pseudocoronata Pres-

cott which comprised upwards of *10–30% of the

assemblage. Mallomonas acaroides Perty was also

present in a few lakes but at low abundances. In

general, Synura taxa were not very abundant, except in

lakes 13C and 15F, where S. sphagnicola reached

*60% of the assemblage (Fig. S2). Synura echinulata

Korshikov reached low abundances in a number of

lakes.

Flux rates of diatoms and chrysophyte scales

Both the concentration and flux rates of scales have

increased with time in all lakes in which they were

present (Fig. 3). The maximum flux rate across the

lakes varied from *26–105 9 104 scales per

cm2 year-1, with an average maximum flux of

*60 9 104 scales per cm2 year-1. The initial timing

of when scales were at sufficient numbers to enumer-

ate varied across the lakes from *1940 to 2003, with

the exception of lake 9D which had sufficient numbers

pre-1900. Many lakes indicate a rapid increase in flux

post-1980 (9D, 15J, 8H, 8D), these lakes also show a

further increase post-2000. The other lakes only

indicate the more recent increase post-2000 (10W,

15F, 7A, 13C), largely because most of these lakes

only began to have sufficient numbers to count

between *1980–1990 (Fig. 3).

Similar to the post-1980 increase in scaled chrys-

ophytes, diatom flux rates increased substantially in

post-1980 sediments in seven of the 10 cores (Fig. 3).

The maximum flux rates across the lakes varied from

*10–125 9 105 valves per cm2 year-1, and up to

440–730 9 105 per cm2 year-1 in lakes 10G and

12D. The average increase in diatom flux within each

lake varied from *1.2 to 2.4 times when comparing

the average of more recent samples to the average of

older samples within each lake. In lakes 9D, 10G and

15J the flux rates in the diatoms are more variable and

lack a pronounced post-1980 increase (Fig. 3). The

diatom flux is significantly correlated to the sedimen-

tation rate in all but lake 9D (Table S1). In general,

diatom numbers were an order of magnitude greater

than the scaled chrysophytes. Interestingly, the two

lakes with the highest diatom flux (with S. pinnata as

the dominant taxa) are those in which the scaled

chrysophytes are rare or absent.

Main direction of variation of diatom assemblages

The first axis of a PCA was used to summarize the

main direction of variation in the diatom species

assemblages in each core (Fig. 3). The minimum and

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maximum PCA scores were used to determine the

overall change within each of the cores, and lakes were

plotted according to increasing change in species

assemblages from lowest (lake 9D, change of 1.2 U) to

highest (lake 10G, change of 3.3 U) (Table 2). The

eigenvalues for axis 1 varied from 0.19 to 0.74, with an

average of 47% of the variance being explained by

PCA axis-1 (Table 2).

Diatom assemblage zones were defined through a

constrained cluster analysis and optimal zonation. The

two techniques defined similar groups in seven of the

ten cores, and the broken-stick determined they were

different from a random assemblage. In those lakes

where the zones varied between the two methods

(lakes 8H, 13C, 10G), grouping was based on the

constrained cluster analysis. The timing of the changes

in species assemblage varied from *1968 to 1997,

with clusters of earlier changes (*1968–1979) for

lakes 9D, 10W, 15J, 8H, 15F; and later changes

(*1990–1997) for lakes 7A, 13C, 12D, 8D, 10G. The

lakes with the latest changes are those with a greater

degree of overall species change as determined by the

overall change in the PCA scores (Table 2), albeit

changes in any particular species were small (Fig. 2).

In many of the lakes, the timing of the change in the

species assemblage, as represented by the PCA axis-1

scores appears to correspond to the increase in diatom

flux rate (Fig. 3). However, a significant correlation

between PCA axis-1 scores and diatom flux is only

present in three lakes (8D, 10W, 15F) (Table 2).

Fig. 3 Summary of results of fossil remains of chrysophyte

scales and diatoms. Flux rates for scales are number 9 104 per

cm2 year-1. Two lakes (12D and 10G) had insufficient

concentration of scales to enumerate. Diatom flux rates are

number of valves 9 105 per cm2 year-1. Main direction of

variation of the diatom assemblages is represented by the

principal components analysis (PCA) axis-1 scores. Diatom

inferences of pH are based on a model developed from 47 lakes

in northern Alberta (Pla & Curtis, 2006). The solid horizontal

lines indicate the two zones based on the diatom assemblages,

the dotted horizontal lines extend these across for ease of

comparison

Hydrobiologia (2013) 720:55–73 63

123

Diatom-inferred pH changes

Diatom-inferred pH changed very little within each

core (Fig. 3), with an average change in pH values of

only 0.22 U (range of 0.11–0.36 U). The majority of

lakes do not have a significant relationship between

pH and the main direction of variation (PCA axis-1

scores), suggesting that any small changes seen in the

diatom species assemblages are not due to changes in

pH. PCA axis-1 scores are only significantly corre-

lated to diatom-inferred pH in two lakes (Table 2). In

lake 10W, the increase in taxa such as C. mediocris

and N. perminuta by *8–9% each, which have lower

pH optima than A. minutissimum (which decreased),

are the main changes driving the small (0.25 U), but

distinct decline in pH starting in the 1970s. PCA axis-1

scores for lake 10W explains *61% of the variation in

species assemblage, so pH or related limnological

variables may be a main driver of this change in

species. In lake 15J, an increase in D. stelligera and S.

pinnata, which have slightly lower pH optima from A.

minutissimum are likely the drivers in the small

(0.36 U) change in pH, which began in the 1940s,

with further declines in the late 1960s. However, PCA

axis-1 scores in lake 15J only explains *19% of the

variation in species assemblages, so pH or related

limnological variables are likely only a small compo-

nent of the changes seen.

Sedimentation rates

Sedimentation rates, based on a single-core per lake,

show clear increases in eight of the 10 lakes, with

slightly more complex patterns in lakes 9D and 15J

(Fig. 4). Minimum rates within each core (typically

near the bottom of the profile) varied between 0.0014

and 0.0047 g cm-2 year-1, with maximum rates

(typically near the top of the core) varying from

0.0046 to 0.018 g cm-2 year-1. Comparing the aver-

age sedimentation rate at the bottom of the profile

(three samples) versus the average rate at the top of the

profile (three samples) the sedimentation rate has

generally increased by 1.5–3.3 times, with a high of

*4.6 times in lake 15F (Fig. 4). Although the timing

of the largest increase in sedimentation rate between

samples within each lake varied tremendously from

Table 2 Largest change in

PCA axis-1 scores based on

the diatom assemblage data

since c. 1900 and the

eigenvalues (k) associated

with the PCA axis-1 scores

The correlations (r) and

associated P values are

provided between PCA

axis-1 scores and diatom-

inferred pH and diatom flux

are also shown (significant

correlations are indicated in

bold)

Lake Number of

samples

PCA axis-1

change

k PCA axis-1 PCA vs diatom-

inferred pH

PCA vs diatom

flux

9D 12 1.16 0.74 0.46

P = 0.13

0.19

P = 0.56

10W 14 1.55 0.61 0.61

P = 0.02

0.74

P = 0.0023

15J 20 1.66 0.19 0.61

P = 0.0044

0.41

P = 0.07

8H 17 1.87 0.42 0.17

P = 0.52

-0.39

P = 0.12

15F 20 2.03 0.32 0.20

P = 0.40

0.78

P < 0.0001

7A 16 2.63 0.42 0.19

P = 0.48

0.18

P = 0.50

13C 20 2.66 0.39 0.20

P = 0.39

-0.36

P = 0.11

12D 17 2.69 0.40 0.46

P = 0.13

0.25

P = 0.49

8D 20 3.10 0.66 0.27

P = 0.26

0.81

P < 0.0001

10G 20 3.27 0.59 0.42

P = 0.06

0.32

P = 0.17

64 Hydrobiologia (2013) 720:55–73

123

*1940 to *2007, most of the lakes (9D, 15J, 8H,

15F, 12D, 10G) begin to increase in sedimentation rate

between *1965 and 1980.

Analysis of siliceous microfossils and metals were

carried out on the top 20 cm of each core. For the

majority of lakes the top 20 cm represented the last

80–100 years. For three lakes (8D, 10G, 15F), overall

sedimentation rates were high, and the top 20 cm only

covered the last *50–70 years. This shorter time

frame still enabled analysis of 20–40 years of pre-

industrial conditions prior to the onset of the large

expansion of the Alberta oil sands beginning in

*1980.

Metal profiles

As a means of summarising the data of trace metal and

other element fluxes for each of the sedimentary

profiles we combined the elements for each lake that

are on the Environmental Protection Agency list of

PPEs (Kelly et al., 2010). The elements combined in

our dataset were: arsenic (As), beryllium (Be),

cadmium (Cd), copper (Cu), chromium (Cr), lead

(Pb), mercury (Hg), nickel (Ni), selenium (Se) and

zinc (Zn). In addition to these PPEs we also have

included vanadium (V), because it is known to be an

atmospheric pollutant associated with the burning of

oil (Norton et al., 1992). Total PPE flux is based on the

total combined concentration multiplied by the sedi-

mentation rate for each sample.

The total flux of PPE to the sediments increased in

all cores (Fig. 4). However, the total flux data is highly

related to sedimentation rates (Table S1). As a means

of distinguishing between natural (watershed and

atmospheric sources) and anthropogenic sources

(atmospheric) of total PPEs, the total PPE flux was

standardized to titanium (Ti). Ti is a commonly used

reference element for standardization (Reimann &

Caritat, 2005). When the total PPE flux is standardized

to Ti, there is little trend in flux in most of the study

lakes. The exceptions to this are for lakes 7A and 12D,

which respectively show increasing and decreasing

trends (Fig. 4).

Another method of distinguishing between natural

and anthropogenic sources of metals and other

elemental pollutants is to subtract background condi-

tions from the total concentration (Renberg, 1986).

However, only three of the lakes (9D, 10W, 7A) reach

relatively stable low background conditions (pre-

1860–1880) in total PPE concentration (Fig. Supple-

mental 3, S3). When examining individual elements,

low background conditions were apparent only in the

concentration of lead and mercury, but not in other

elements, such as copper, zinc and vanadium (Figs.

Fig. 4 Summary of sedimentation and flux rates of total

priority pollutants (PPE). See the Results ‘‘Metal profiles’’

section for the list of elements combined in the total PPE.

Sedimentation rates of the single-core per lake are in

g cm-2 year-1. Total PPE flux and total PPE flux standardized

to titanium (Ti) are in lg cm-2 year-1

Hydrobiologia (2013) 720:55–73 65

123

Supplemental 4–8, S4–S8). Two of the other longer

lake records (8H, 12D) nearly reached background

conditions in lead and mercury, whereas the shorter

records (post 1920) do not.

The concentration and flux rate of individual

elements varied tremendously between lakes (Figs.

S4–S8). There was similarity in the profiles of lead

concentration in the longer lake records, with increas-

ing concentrations post *1880 (post *1920 in lake

7A), reaching highs at *1980–1990s, and then

decline (Fig. S4). This pattern was similar for mercury

(Fig. S5). For elements such as copper, zinc and

vanadium, variability was high with little similarity

between lakes (Figs. S5–S8).

Discussion

Lack of acidification

One of the main objectives of this study was to

determine whether any recent lake acidification

occurred in the northwest Saskatchewan study lakes,

which might be associated with the nearby develop-

ment and expansion of Athabasca oil sands in Alberta.

Only two lakes indicated any signs of a consistent

decline in diatom-inferred pH that was significantly

related to the main direction of variation in species

assemblages, as measured by PCA axis-1 scores.

Comparison of inferences and the main direction of

variation enabled an evaluation of the relevance of the

inferences in relationship to other potential drivers of

species changes. The degree to which the change in

species assemblage in these two lakes may be related

to the pH change (or other limnological variables

affected by a pH change) varied tremendously, as is

evident from the amount of variation explained by the

PCA axis-1 scores of 61% for lake 10W and 19% for

lake 15J (Fig. 2). Lake 10W is one of the closest sites

to the oil sands mining operations (Fig. 1), and

although diatom species changes are small, inferred

pH does indicate slight acidification since the mid-

1970s, potentially related to atmospheric deposition

associated with development. On the other hand, lake

15J is one of our furthest sites, but the low variation

explained in the PCA analysis, suggests that even with

a significant relationship between inferred pH and

PCA axis-1 scores, pH or related variables are likely

not the only drivers of the small changes in the diatom

assemblages. For both lakes, the decline, although

distinct, was relatively small (0.25–0.36 pH units), and

within the errors of the predictive models.

Although one of the closest lakes to the oil sands

mining operations, lake 10W does not stand out as being

different in our study sites in terms of size or chemistry.

However, when compared to the other closest study

lakes (8D, 10G), lake 10W clearly has lower alkalinity

and specific conductance, as well as slightly lower

chlorophyll a and total phosphorus (Table 1). The lower

buffering capacity (\100 leq l-1) for lake 10W in

conjunction with its location would make it more

susceptible to acidification than lakes 8D and 10G, our

two most buffered sites ([100 leq l-1). In the northeast

USA, only lakes with extremely low buffering capacity

(ANC \ 50 leq l-1) were susceptible to chronic acid-

ification (Sullivan et al., 1990; Cumming et al., 1994).

Although lakes 8D and 10G do not indicate any signs of

acidification, these two lakes have the largest change in

species assemblages, albeit small (i.e. taxa present

*100 years ago are similar to today, but % abundances

changed), as indicated by the change in PCA axis-1

score values (Table 2). In both of these lakes, there is an

increase in the planktonic D. stelligera (and D.

pseudostelligera in lake 8D) and a decline in the more

benthic S. pinnata, suggesting a shift within the lake

ecosystem. However, these shifts occurred in the late

1990s, much later than the change at lake 10W,

suggesting that there were likely very different drivers

of the changes observed in these lakes.

A lack of any recent apparent acidification in most

of our study lakes is similar to the findings which

examined Alberta lakes within the Athabasca region

(Hazenwinkel et al., 2008; Curtis et al., 2010). In the

20 combined lakes examined in these two studies, only

one showed any strong evidence of acidification,

whereas a number of lakes indicate signs of nutrient

enrichment. The one lake that showed a decline in pH,

was the smallest (11 ha), one of the shallowest and has

the shortest hydrological residence time (Curtis et al.,

2010). Along with its low alkalinity, these other

characteristics likely made it more susceptible to

atmospheric acid deposition (Curtis et al., 2010). In

general, the size of our study lakes was smaller than

the eight studied by Hazenwinkel et al. (2008, size

range = 103–4,322 ha, only our two largest fall into

the lower size range), and the 12 studied by Curtis

et al. (2010, size range = 11–955 ha, our two largest

fall in the upper range, whereas the rest are smaller

66 Hydrobiologia (2013) 720:55–73

123

than their third largest at 55 ha). Our lakes also tend to

have lower alkalinity, nutrients and DOC than the

Hazenwinkel et al. (2008) lakes, but fall within the

range of lake chemistries examined by Curtis et al.

(2010). Thus, our study which extends the number of

small, low alkalinity, low nutrient and low-DOC lakes

(those felt to be the most susceptible to acid deposition

in this region), suggests that any increases in atmo-

spheric acid deposition to these lakes has not resulted

in chronic acidification, with the potential exception of

lake 10W. Our study sites are further downwind of the

oil sands mining operations than the other two lake

studies and outside of the known current range of high

deposition of PAHs and other airborne pollutants

found in snowpack and tributaries within *50 km of

oil sands mining facilities (Kelly et al., 2009, 2010).

The slight decline in pH of lake 10W, one of our

closest sites to oil sands mining operations (but

[50 km) provides some evidence that emission of

sulphur and nitrogen compounds associated with this

mining may have caused this small decline in pH.

Atmospheric transport of metals

Profiles of metals in sediments have been used to help

determine if emissions from industrial development

have increased (Baron et al., 1986; Norton et al.,

1992). A large suite of metals have been found in

atmospheric deposition (Galloway et al., 1982),

including all of the PPE pollutants examined in this

study. It is well established that many of these

pollutants, particularly mercury, which has been

widely studied around the northern Hemisphere, can

be deposited broadly through long-range atmospheric

transport and deposition (Fitzgerald et al., 1998).

Atmospheric deposition of Pb, Cu, Zn and more

recently V have been shown to increase in regions

downwind of fossil-fuel emissions, particularly in

eastern United States (Baron et al., 1986). Anthropo-

genic releases of heavy metals, such as Cu, Pb and Zn,

to the atmosphere have outpaced any of the natural

sources (Nriagu, 1979, 1990, 1996). Although many of

these pollutants can be transported great distances, the

highest concentrations are typically found closer to the

source (Rognerud & Fjeld, 2001; Skjelkvale et al.,

2001; Augustsson et al., 2010). Much of this evidence

has come from the analysis of lake sediment records.

Determining if emissions have increased from

sedimentary evidence in part relies on defining

background conditions that originate from natural

sources. Defining background conditions can be

difficult in some regions, particularly remote regions

where natural processes may outpace the relative

contribution from atmospheric sources (Augustsson

et al., 2010). Furthermore, ancient cultures have also

been shown to produce significant quantities of heavy

metals (Pb, Cu, Zn, Hg) that resulted in substantial

emission to the atmosphere (Nriagu, 1996). Bindler

et al. (2011) suggest because of this ancient contribu-

tion of contaminants to the atmosphere that back-

ground conditions should be based on sediments

[3,000 years old. However, in many studies this is

impractical, and depending on the question and region

may not be necessary. Assessment of recent increases

on a number of sites within the same region provides a

means of determining whether there are consistent

trends across the sites. A design of multiple sites also

enables the ability to determine the degree of between

lake variability, and determine the degree of variabil-

ity that may be due to natural catchment processes

versus atmospheric influences both pre and post

disturbance.

Our study indicated that there was a large degree of

variability between the lakes in terms of both concen-

tration and flux rates of total priority metal pollutants.

This was also the case when examining individual

pollutants. The sedimentary profiles from our longer

temporal records suggested that background condi-

tions were reached pre-1860–1880. Low stable back-

ground conditions were particularly apparent for both

the concentration and accumulation (flux) of Pb and

Hg (Figs. S4–S5). Whereas, the profiles of elements

such as Cu, Zn and V (Figs. S6–S8) indicate a higher

degree of natural variability in this region with no

distinct low and stable background conditions. Deci-

phering recent changes associated with anthropogenic

emissions from natural processes may be difficult for

some pollutants in this region, where the natural

variability is high and atmospheric deposition is

relatively low.

Another source of difficulty is changes in the rate of

sediment deposited into the lake can highly influence

the concentration of metals in the sediments (e.g.

Renberg, 1986; Rognerud et al., 1998). If sedimenta-

tion rates increase significantly then the concentration

of metals can be diluted regardless of any changes in

atmospheric deposition. Our study indicated that

sedimentation rates have significantly increased in

Hydrobiologia (2013) 720:55–73 67

123

all of our lakes, thus recent declines in the concentra-

tion of total PPE in the majority of our lakes (Fig. S3)

is likely due to increased sedimentation from allo-

chthonous and/or autochthonous sources such as

increased algal production. Because flux rates are

based on concentration times the sedimentation rate,

the increases in the total PPE flux in many of our study

lakes (but decline in concentration) are likely heavily

tied to the increase in sedimentation rather than from

increases in atmospheric deposition. As a conse-

quence, we examined both concentration and accu-

mulation (flux) rates of pollutants as suggested by

Renberg (1986), so that a more robust interpretation

could be made.

Another method to assess changes in metals is to

standardize the concentrations to a reference element,

such as aluminium (Al) or titanium (Ti) (Reimann &

Caritat, 2005). When we standardized to Ti, our data

indicated that the majority of lakes did not show any

recent increases. This provides some evidence that

there has been little change in deposition into these

lakes. However, some studies suggest that the use of a

reference material for normalization of pollutant flux

rates may not always reliably demonstrate anthropo-

genic influences from natural processes due to the

potential variability of the reference material from

atmospheric and watershed processes (Norton et al.,

1992; Reimann & Caritat, 2005; Augustsson et al.,

2010). Nonetheless, our conclusion of little change in

the rate of priority pollutants to these lakes since the

increased development of the oil sands is similar to the

finding from the generally low concentrations of PAHs

in these lake sediments, with no discernible trends with

the exception of a noticeable increase in PAH concen-

tration and flux in lake 8D, one of the closest sites to oil

sands mining operations (Ahad et al., 2011). The most

abundant PAHs measured was perylene, which is a

product of microbial degradation of organic matter

during sediment diagenesis (Wakeham et al., 1980;

Grice et al, 2009) and retene, which is a marker for

coniferous wood combustion (Ramdahl, 1983: Benner

et al., 1995). Similarly, analysis of d15N down core

indicated no consistent discernible trends associated

with the onset of bitumen mining (Ahad et al., 2011).

Other potential drivers of change

Rates of deposition of sedimentary materials into lakes

are influenced by both allochthonous and autochthonous

processes, which in turn can be influenced by changes in

climate and/or landscape processes. Annual mean

temperature in Fort McMurray has generally increased

from an average of -1.1�C between 1916 and 1940, to

an average of 1.3�C from 1987 to 2011, with the

majority of years above the record average starting in the

1970s (Fig. 5). Several periods of below-average pre-

cipitation occurred, with the 1940s to early 1950s and

since the late 1990s being most distinct (Fig. 5). These

climatic patterns are consistent with regions of the

Canadian prairies and boreal transition zones in which

many sites indicate a rise of *1–4�C, with much of the

increase in temperature since 1970, and with recent

declines in precipitation (Schindler & Donahue, 2006).

Although there is certainly not a consistent pattern in the

timing of the onset of increased sedimentation rates,

many of the lakes do indicate an increase beginning in

the 1970s and 1980s, with others more recently in the

1990s and 2000s. The correspondence of similarity in

timing of events in the lakes, does suggest a similar

forcing driving the changes. In general, more productive

lakes have higher sedimentation rates (Umbanhowar

et al., 2011). Warming may be leading to a higher influx

of sediments to the lake (allochthonous inputs) or may

be leading to increased algal production (autochthonous

inputs) or a combination of the two. These processes

may in part be the drivers behind the increased

sedimentation rates seen in the 1970s and 1980s,

however the later changes (post 1990s) are also

associated with a sharp decline in precipitation, sug-

gesting there may be some other mechanism for the

continuing increase in sedimentation rates. While

evaluation for potential linkages between climate and

increased rates of sedimentation and diatom production

warrants further study, similarity of trends in these

variables across multiple watersheds suggests a large

spatial-scale forcing such as climate.

Increased sedimentation rates may also be in part an

artefact of the CRS dating model, which may overes-

timate these rates at the top of cores with flattened210Pb profiles, particularly for lakes 15J, 15F, 8D, 10G

(Fig. S1). However, this pattern was not consistent

across all of our lakes, suggesting this mechanism is

not the sole driver of the changes seen in sedimenta-

tion rates. Another complication is that inferring

sedimentation rates from a single core can be prob-

lematic due to differences in deposition across the lake

basin, particularly sediment focusing in the central

basin (Engstrom & Wright, 1984). Thus while the

68 Hydrobiologia (2013) 720:55–73

123

estimates from a single core give us some idea of

changes in sedimentation rates, ideally sedimentation

rates should be based on multiple cores across the

basin.

There are also two apparent different times of

change in the diatom species assemblages, with half of

the lakes indicating change in the 1970s and 1980s,

and the other half in the 1990s and 2000s; however,

these groups do not consistently correspond to the

timing of sedimentation changes within each lake.

Although clustering techniques indicate changes in the

diatom assemblages that are recognizably different

from a randomized dataset, individual species changes

are small in all lakes.

Diatom flux increased in most lakes, suggesting that

the production of diatoms has increased. Although

these increases are related to the increase in sedimen-

tation rates, there is fairly good correspondence

between the timing of the increase in flux with the

timing of the small change in species assemblage

(Fig. 3), suggesting that changes in the diatom flux are

likely not just solely an artefact of increased sedimen-

tation. Climatic warming can have a number of effects

on the physical, chemical and biological properties of

lakes (Adrian et al., 2009). Properties such as water

temperature, ice phenology and chemical composi-

tion, if measured through time, can provide a diverse

array of information on changes potentially associated

with climate. Sediments in the lakes archive this

information, which can be used to help decipher

changes on yearly to millennial scales (Williamson

et al., 2009). Increased temperatures generally leads to

increased growth, if the organisms of interest are not

limited by other resources, and can also lead to

compositional changes in species assemblages

(Adrian et al., 2009). Increasing warmth has been

shown to result in substantial changes to the timing of

ice out and length of stratification, which in turn can

result in changes to algal production and composition

(Winder & Schindler, 2004; Winder & Sommer,

2012). Data from the analysis of alkanes and C/N

ratios suggest that there was an increasingly larger

component of algal-derived organic matter in many of

our study lakes (Ahad et al., 2011), providing further

evidence of increased algal production (e.g. Basascio

& Bradley, 2012). Such evidence suggests that climate

may be an important factor of the changes seen both in

the flux and species composition of the diatoms.

Although our focus here has been on temperature

effects, changes in precipitation could result in

changes within the watershed. Sharp declines in

precipitation occurred in the Fort McMurray record,

particularly in the 1940s, however there is little

evidence for change within the lake around this time.

Climate was also surmised to be a potential driver of

the diatom changes seen in the Alberta lakes near Fort

McMurray (Hazenwinkel et al., 2008; Curtis et al.,

2010). While increased algal production could also be

the result of increased nutrient loading from internal

lake dynamics, forest fires or atmospheric deposition

(Hazenwinkel et al., 2008; Curtis et al., 2010; Greaver

et al., 2012), individual diatom species changes in our

study lakes were small and none were clearly indic-

ative of nutrient changes.

There was a diversity of changes in diatom species

composition across the lakes. In a number of lakes

small increases in Discostella species occurred in both

the c. 1970 changing lakes (15J), as well as in the lakes

with the c. 1990 temporal changes (7A, 8D, 10G).

Recent increases in Discostella species have been

surmised to be indicative of increased stratification as

a result of lake warming (Ruhland et al., 2008).

However, this pattern was not consistent in our study

lakes. The other lakes with Discostella, D. stelligera

either declined (9D) or remained at relatively low

percentages (15F, 12D). The decline of D. stelligera in

lake 9D beginning in the early 1980s corresponded

with the timing of the decrease in diatom flux. Other

Fig. 5 Mean annual temperature (�C) and annual precipitation

(mm) for Fort McMurray, Alberta. The horizontal lines are

depicting the mean of the records. Data are from the Adjusted

and Homogenized Canadian Climate Data archive (http://ec.gc.

ca/dccha-ahccd)

Hydrobiologia (2013) 720:55–73 69

123

changes in the planktonic composition of the lakes,

albeit a smaller component of the entire assemblage,

may also be indicative of changes in stratification or

other related limnological variables. Fragilaria tenera

increased slightly in the most recent samples of lake

10W, this is the only planktonic taxa in this record,

suggesting that the lake environment has sufficiently

changed to enable a planktonic component to the

assemblage. Several lake records (8H, 13C, 12D) were

comprised of a small percentage of Aulacoseira taxa,

which are typically planktonic or tychoplanktonic. In

all of these cases, Aulacoseira decreased towards the

top. The larger Aulacoseira taxa (A. subarctica, A.

ambigua), which can form long chains, are planktonic

and need turbulent conditions to remain in the photic

zone, suggesting stratification may have increased

with the decline of these taxa (Ruhland et al., 2008).

Whereas, the smaller Aulacoseira (A. lacustris, A.

peraglabra), have been associated with mid-depth

benthic habitats in Ontario boreal lakes (Kingsbury

et al., 2012), and thus have a more tychoplanktonic life

strategy. Thus whether the decline of these Aulacose-

ira taxa are linked to changes in stratification are less

clear.

Although there were marked changes in the flux of

scaled chrysophytes, there were very few changes in

terms of the species assemblages. The majority of lakes

were dominated by M. crassisquama throughout the

cores. This taxon is indicative of circumneutral to

alkaline lakes. Thus the scaled chryosphytes also do

not indicate any marked changes in the pH of the study

lakes. Only two lakes indicated any significant change

in the species assemblages. Lake 15F had a sharp

increase in S. sphagnicola in the upper sediments (post-

2008), whereas lake 13C indicated a decline in this

same species. S. sphagnicola does have a lower pH

optimum (Cumming et al., 1992), however increases in

the colonial Synura have also been linked to increased

stratification amongst other factors (Hyatt et al., 2010).

Although our lakes indicate some evidence of

increased production, the species changes in our lakes,

are much more subtle than in some of the Alberta lakes

which showed pronounced increases in Stephanodis-

cus minutulus (Hazenwinkel et al., 2008) and Aste-

rionella formosa (Curtis et al., 2010), taxa which are

indicative of higher nutrient conditions. Increases in A.

formosa and Fragilaria crotonensis have been linked

to increased nitrogen deposition in alpine lakes in

western US (Saros et al., 2011), both of these taxa are

rare or absent in our study lakes. However, the

increase in scaled chrysophytes in many of our study

lakes from very low concentrations or absence prior to

ca. 1940–1960 to larger populations recently (post-

1980–2000) indicates that there have been fundamen-

tal changes in the lakes enabling the expansion of this

planktonic group of organisms.

Conclusions

Evidence from paleolimnogical data is consistent with

analysis of lake-water chemistry (Scott et al., 2010)

and process-based models (Whitfield et al., 2010)

which all indicate that deposition of strong acids has

not resulted in chronic acidification in northern

Saskatchewan. The one lake that slightly acidified

and had the main direction of variation highly related

to pH is one of the closest sites to the oil sands

development. The two other closest sites have higher

alkalinity and thus higher buffering capacity.

Although subtle, other diatom species changes suggest

that climate may be having an influence on the diatom

assemblage potentially through changes in the length

of the ice-free season, strength of stratification, and

other related variables. The diatom species changes do

not suggest increases in nutrients that could occur with

climate and internal phosphorus loading, rather the

increase in overall diatom flux may be a signal of

increased production due to warming. Increased

production could also be the result of increased

nitrogen deposition; however, the specific diatom

species changes do not clearly support this.

There were no consistent discernible trends in total

or individual PPEs measured in this study. Concentra-

tion of lead and mercury has increased in many of the

lakes since ca. 1860, but decreased in recent sediments,

likely as the result of increased sedimentation rates.

There were no discernible patterns in concentration

across lakes for other pollutants such as copper, zinc

and vanadium, indicating that at the low concentrations

of these elements it is difficult to disentangle anthro-

pogenic sources from background natural sources in

this region. Although the flux of total PPE suggested

increasing trend across all of our lakes, when these

values were standardized to Ti the consistent increas-

ing trends were no longer discernible.

While a lack of widespread acidification, or

increased concentrations of priority pollutants,

70 Hydrobiologia (2013) 720:55–73

123

indicates there has been no large discernible impact

from the expanding oil sands development to date, the

continued expansion may lead to different results, and

lakes in the region need to be further researched and

monitored to ensure that expansion does not begin to

have larger impacts. The changes seen in both the

diatoms and scaled chrysophytes indicate that these

lakes are not in a stable regime, and that climate,

internal lake dynamics or potentially atmospheric

deposition may be playing a role.

Acknowledgments This project was funded by the

Saskatchewan Ministry of the Environment. We would like to

thank Kenneth Scott for assistance with lake selection and

fieldwork, Steve Wilke for assistance in the field, and Moumita

Karkamar for assistance in the lab. We are also thankful to Oil

Sands Quest Inc. for use of their facilities while undertaking the

fieldwork associated with this project.

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