Mercury concentrations in amphipods and fish of the Saint Lawrence River (Canada) are unrelated to concentrations of legacy mercury in sediments

Science of the Total Environment 494–495 (2014) 218–228

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Mercury concentrations in amphipods and fish of the Saint Lawrence River (Canada) are unrelated to concentrations of legacy mercury in sediments Peter V. Hodson a,⁎, Kristin Norris a,1, Michelle Berquist a,2, Linda M. Campbell a,3, Jeffrey J. Ridal b a b

Department of Biology, Queen's University, Kingston, Ontario, Canada Saint Lawrence River Institute of Environmental Sciences, Cornwall, Ontario, Canada

H I G H L I G H T S • • • • •

Sediments of L. St. Francis are contaminated with Hg from past industrial activity. However, Hg in resident aquatic biota does not reflect Hg distribution in sediments. Hg in biota may be derived from current loadings, likely from tributaries. Managing current sources of Hg will be more beneficial than remediating sediments. Managing the risks of Hg to fish consumers requires spatially intensive sampling.

a r t i c l e

i n f o

Article history: Received 14 April 2014 Received in revised form 27 June 2014 Accepted 28 June 2014 Available online xxxx Editor: Mae Sexauer Gustin Keywords: Mercury Sediments Amphipods Fish Heterogeneity Risk assessment

a b s t r a c t Past industrial activity at Cornwall, Ontario, Canada has contaminated Lake Saint Francis, a fluvial lake on the Saint Lawrence River, with mercury (Hg). A spatial survey of Hg concentrations in sediments, amphipods, and yellow perch (Perca flavescens) in 2008 inferred current sources of Hg to the lake and spatial variations in risks to human consumers. Patterns of total and methyl Hg concentrations in sediment reflected upstream inputs, declining concentrations downstream, and highest concentrations at north shore sites near industrial sources; concentrations were lowest on the south shore because river currents limit north–south advective exchange. Surprisingly, concentrations of total or methyl Hg in sediments and pore water were unrelated to concentrations in amphipods and yellow perch. Concentrations in biota, and risks to consumers of fish, were highest at north shore sites near tributaries, and not at the most contaminated industrial sites. These results suggest that ‘legacy’ Hg in surficial sediments is not bioavailable to aquatic biota; tributaries and atmospheric deposition are possible sources of bioavailable Hg; and that sediment remediation would not resolve issues of Hg in fish. Fish consumption advisories for the entire lake based on single samples of fish could over- or under-protect consumers, depending on sampling location. To understand the actual risk to fish consumers for a large and complex lake system with multiple sources of Hg, more intensive sampling is needed to assess the spatial distribution of risk. © 2014 Elsevier B.V. All rights reserved.

1. Introduction Lake Saint Francis (LSF), a fluvial lake on the Saint Lawrence River (SLR), is situated between the upstream cities of Cornwall, Ontario (ON), Canada and Massena, New York, USA, and a downstream ⁎ Corresponding author at: School of Environmental Studies, Queen's University, Biosciences Complex, 116 Barrie St, Kingston, Ontario K7L 3N6, Canada. Tel.: +1 11 613 533 6129; fax: +1 11 613 533 6090. E-mail address: [email protected] (P.V. Hodson). 1 Current Address: 36 Foxmeadow Drive, Stoney Creek, Ontario, Canada. 2 Current Address: 44 - 1576 Bathurst St., Toronto, ON, M5P3H3, Canada. 3 Current Address: Environmental Science, Saint Mary's University, Halifax, Nova Scotia, Canada.

http://dx.doi.org/10.1016/j.scitotenv.2014.06.137 0048-9697/© 2014 Elsevier B.V. All rights reserved.

hydroelectric dam at Beauharnois, Quebec, (Qc), Canada (Fig. 1). In 1986, LSF was declared an Area of Concern (AOC) by the International Joint Commission which oversees the Canada–US Agreement on Great Lakes Water Quality (1978). Among the primary concerns was the contamination of fish by mercury (Hg) from industrial activities at Cornwall, ON, including a chlor-alkali plant, a pulp and paper mill, and a rayon mill. To manage the risks of adverse effects on human health and aquatic ecosystems, Federal, State, Provincial, Tribal and First Nation governments initiated public health advisories to limit consumption of sports fish due to mercury (Hg) and PCB contamination. To resolve the Hg issue, they developed a Remedial Action Plan to identify and control sources, to remediate industrial sites, and to restore the beneficial uses impaired by contamination (Ridal et al., 2010).

P.V. Hodson et al. / Science of the Total Environment 494–495 (2014) 218–228

Length - 50 km Width - 4.7 km Surface area – 272 km2 Mean depth - 6 m Volume - 2.8 km3 Mean discharge – 7,500 m3/s

16 15 14 12 10

11

17 19

8

22

2

23

Cornwall

6

North shore

3 1

20 21

24

5 4

18

13

9

7

219

25

South shore

26 27

Fig. 1. Sample sites and bathymetry in L. St. Francis, Saint Lawrence River www.ec.gc.ca/stl/default.asp?lang=En&n=09C5A944-1#stf). The River flows from SW to NE, and a hydroelectric dam is just upstream of Cornwall, off the map. A downstream dam at Beauharnois, Qc, is just off the map to the northeast. Numbers and circles refer to sample sites.

1979 Cornwall

1989 Cornwall

1999 Cornwall

2008 Cornwall

Massena 10 km

THg Concentration (µg/kg) 0.0

0.1

0.2

0.3

0.5

Fig. 2. Total mercury concentrations (μg/kg dw) in surficial sediment (top 5-cm) of L. St. Francis, Saint Lawrence River (adapted from Pelletier, 2010, with permission). The contour map infers the concentrations from sample sites on the north shore (red) and south shore (yellow).

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P.V. Hodson et al. / Science of the Total Environment 494–495 (2014) 218–228

The concentrations of Hg in surface sediments downstream from Cornwall are higher on the north shore than the south (Fig. 2) and decrease from upstream to downstream, suggesting that industries at Cornwall were the primary source of a net downstream distribution of Hg. Dated sediment cores demonstrated Hg loadings to sediments prior to 1970 of N 410 g/d, declining to about 20 g/d by 1995 after emissions controls and closure of industries (Delongchamp et al., 2009). As a consequence, there was a steady 56% overall decline in Hg concentrations of surface sediments from 1979 to 2008 (Fig. 2) as less contaminated sediments accumulated over more contaminated. Nevertheless, concentrations of methylmercury (MeHg), the form of Hg that most strongly bioaccumulates, were 190-fold higher in recent surface sediments of the SLR at Cornwall, compared to a reference site (Delongchamp et al., 2010). Given the potential to disperse contaminated sediments in shallow waters, administrative controls on development along the Cornwall waterfront were implemented in 2006 (Environment Canada; http:// www.rrca.on.ca/view.php?id=40). The intent was to avoid activities causing sediment re-suspension and increased Hg methylation, and to facilitate the burial of Hg-contaminated sediments by the gradual accumulation of less contaminated sediments. Emission controls were supported by research along the Cornwall waterfront showing spatial and temporal trends in contamination of fish related to industrial and municipal outfalls (e.g., Fowlie et al., 2008; Choy et al., 2008; Ridal et al., 2010). Monitoring of Hg in fish in the SLR by the Ontario Government has been at a regional scale, based on ‘river blocks’, i.e., long stretches of rivers with similar characteristics, with three upstream blocks between L. Ontario and Cornwall separated from LSF by a dam at Cornwall-Massena. Since 1975, Hg concentrations in northern pike (Esox lucius), smallmouth bass (Micropterus salmoides), walleye (Sander vitreus), and yellow perch (Perca flavescens) declined by 60 to 65% throughout the SLR, reflecting basin-wide responses to Hg control, including reduced atmospheric deposition (Neff et al., 2013). In LSF, these changes were also associated with controls of industrial emissions at Cornwall. Contamination of LSF fish with Hg is consistent with established models of Hg cycling in aquatic environments (Merritt and Amirbahman, 2009; Jaglal, 2010). According to these models, the precipitation of inorganic Hg to sediments in association with organic particulates contributes to a pool of bioavailable Hg that can be methylated by sulfate-reducing bacteria to MeHg in the surface layers of sediments (Skyllberg et al., 2007; Kim et al., 2008). The exposure of benthic communities to MeHg, either by contamination of pore waters or by food web transfer to invertebrate grazers, initiates a transfer of MeHg through aquatic food webs, with progressively higher concentrations of Hg, and higher proportions of MeHg, at each trophic level (Rolfhus et al., 2011). At higher trophic levels, fish accumulate the most MeHg, which represents a threat to fish and to wildlife and human consumers when tissue concentrations exceed thresholds for neurotoxicity (Gilbertson, 2009; Scheuhammer et al., 2008; Sandheinrich and Wiener, 2011). The de-listing target for the St. Lawrence River (Cornwall) AOC will be achieved when Hg concentrations in fish of LSF are equivalent to those in upstream river blocks, restrictions on fish consumption are the same as or fewer than in upstream blocks, or when concentrations in fish fall below a consumption guideline of 260 μg/kg wet weight (ww) recommended by the Ontario Ministry of the Environment (OME, 2011) for sensitive consumers. However, the overall rate of decline of Hg concentrations in some species of fish appears slower than the decline in Hg concentrations in sediments (compare Pelletier, 2010 to Neff et al., 2013); in some cases, concentrations in pike, bass, walleye, and perch have increased slightly since 1990 (Neff et al., 2013). While the geometric mean concentrations in perch muscle currently fall below the guideline, variance remains high, with concentrations in individual fish of all sizes in 2010–11 ranging from 100 to 1100 μg/kg ww. This implies a longer time than desired to achieve de-

listing targets, and on-going risks to consumers of fish from LSF. More importantly, it challenges the idea that management focussed on legacy Hg contamination of sediments will resolve the issue of Hg in fish, and that the pattern and extent of sediment contamination provides a reasonable basis for identifying potential risks to consumers of fish. One possible cause of the discrepancy between Hg trends in fish and sediment may be differences in sampling intensity. Mercury concentrations in juvenile perch (1+ to 3+ y; Fowlie et al., 2008) and young-ofthe-year spottail shiners (Notropis hudsonius; Choy et al., 2008) in the SLR at Cornwall varied considerably over spatial scales as small as 500 m. This high spatial variance in Hg content of relatively sessile species implies that the spatial heterogeneity in Hg risks would not be well represented by regional scale monitoring of fish. Assumptions that fish captured at one or two locations represent the risk to consumers across the entire lake may under-state risks in some locations and over-state risk in others. In 2008, we initiated a detailed spatial survey throughout LSF of Hg concentrations in sediments, yellow perch, and amphipods to: infer current sources of Hg to the food web of LSF; assess which measures of Hg in sediment and biota of LSF best indicate the potential ecological and human health risks of Hg concentrations in fish; determine whether one sample of fish within the LSF river block is sufficient to estimate the risks to consumers of fish throughout LSF; and recommend the most effective strategy for monitoring the ecological and human health risks of Hg contamination of fish. 2. Materials and methods 2.1. Study sites Lake St. Francis is 50 km long, up to 4.7 km wide, and relatively shallow (b10 m) except for a shipping channel with a depth of up to 25 m (www.ec.gc.ca/stl/default.asp?lang=En&n=09C5A944-1#stf). Annual flows through LSF average 7500 m3/s and are controlled by the Moses-Saunders hydroelectric dam just upstream of Cornwall and Massena. The water source is predominantly Lake Ontario, with tributary flows comprising 2% of the total. Tributaries on the north shore drain agricultural lands and wetlands, while those on the south shore originate in the forested hills of the Adirondack National Park, NY. Sedimentation and sediment characteristics of LSF are described in detail by Pelletier and LePage (2003). Twenty-seven sites from around LSF were sampled from July to September, 2008: 2 at Cornwall, 12 along the north shore, and 13 along the south shore (Fig. 1). Sediments at Cornwall sites were in close proximity to past industrial sources and highly contaminated with Hg (Fig. 2; Pelletier, 2010). Sediment Hg concentrations at north shore sites were higher than background but declined with distance from known sources. South shore sites had little sediment Hg contamination and were separated from north shore sites by a shipping channel in which strong currents limit north–south advective exchange. Previous reports of sediment THg concentrations in LSF (Pelletier and LePage, 2003) guided the choice of sampling sites to ensure a wide range of Hg concentrations among sites, including tributary mouths. Most sites were in shallow water (b 3 m) close to shore, where sediments would reflect Hg inputs from land-based activities, and where methylation rates should be highest (Munthe et al., 2007). 2.2. Sampling 2.2.1. Sediment and pore water Five sediment samples were taken with a gravity corer at each of the 27 sites along 250–300 m transects parallel to shore, with at least 25 m between each. Core tubes were pre-cleaned with a soft-bristle brush and laboratory grade detergent and rinsed with copious amounts of tap water, 10% HCl, and reverse osmosis (RO) water. Clean cores were covered with clear plastic bags during transport to each site. Cores were

P.V. Hodson et al. / Science of the Total Environment 494–495 (2014) 218–228

taken to the laboratory immediately and extruded slowly to decant all but a thin layer of surface water to prevent oxidation of the sediment. The upper 5 cm of sediment was extruded and transferred rapidly to a plastic bag purged with nitrogen before and after sediment transfer, and bagged samples were placed immediately in a sealed glove box filled with nitrogen gas. Each sample was mixed manually, divided among four 50-ml Falcon tubes, capped, and centrifuged at 3500 rpm for 10 min to obtain pore water, which was decanted under nitrogen. The four subsamples were combined into one tube, centrifuged at 3500 rpm for 10 min, filtered with a 0.45-μm polyether sulfone membrane syringe filter, preserved with trace metal grade HCl, and stored at 4 °C until MeHg analysis. The centrifuged sediments were frozen for THg and MeHg analysis, and residual uncentrifuged sediment was refrigerated for assays of percent loss on ignition (%LOI) and sediment grain size. 2.2.2. Amphipods Amphipods (predominantly Echinogammarus ischnus) were collected with rock basket traps 20 × 20 × 20 cm, containing pebbles 2 cm in diameter (Razavi et al., 2013). Five baskets were deployed at each site where sediment cores were taken, anchored for 7 d, retrieved, and placed immediately in river water. Amphipods were separated from the pebbles, placed on acid-washed tin-foil kept on ice, counted, weighed, and stored frozen (−20 °C) until THg and MeHg analysis. 2.2.3. Yellow perch Gill nets of three, 15 m panels of 5 cm stretch mesh were set at each site for 1–4 h. If 20 yellow perch were not caught, the net was re-set on a different day. Weight, and total, standard, and fork length were measured (Table 1), and scales removed near the head of each fish for aging by plastic impression (Casselman, 1987). Skinless, boneless, white muscle was removed from above the lateral line on both left and right sides of each perch. Each piece was weighed, wrapped in acid-washed foil, and stored frozen until THg analysis, assuming that MeHg comprised more than 90% of THg in fish (Bloom, 1992; Gonzalez et al., 2006). After each fish, dissection tools were washed with 10% nitric acid and rinsed three times with double deionized water. 2.3. Mercury analysis Before analysis, sediment, amphipod, and yellow perch samples were dried and ground. Total mercury (THg) in sediment, amphipods, and yellow perch was measured at the Analytical Services Unit of Queen's University with a Milestone Direct Mercury Analyzer 80 (DMA-80) at a detection limit (DL) of 4.0 μg/kg. Certified reference materials were analyzed in duplicate at the beginning of each run (MESS-3 and TORT-2; National Research Council of Canada, Ottawa, ON). Mean (±SD) THg concentrations were 86 ± 8 μg/kg (n = 60; MESS-3) and 297 ± 36 μg/kg (n = 62; TORT-2), close to expected values of 91 ± 9 μg/kg for MESS-3 and 270 ± 60 μg/kg for TORT-2; recoveries were 95% and 110%, respectively.

221

The repeatability of analyses was assessed by the coefficient of variation (CV) for standard reference materials, and from duplicate measures of unknowns, following a relative standard deviation (RSD) method of the Canadian Association for Laboratory Accreditation (http://www.cala.ca/P19_CALA_Unce_Pol.pdf). The CVs for all MESS-3 and TORT-2 analyses were 9.5% (n = 60) and 12.3% (n = 61), respectively. Every tenth unknown was a duplicate, with an RSD of 8% for sediment (n = 10 pairs), 2% for amphipods (n = 2 pairs), and 8% for yellow perch (n = 32 pairs). Methylmercury concentrations were analyzed at the University of Ottawa using a Hewlett Packard® capillary gas chromatograph coupled to an atomic fluorescence spectrometer (GC-AFS) with a DL of 0.02 ng/L for pore water and 0.016 μg/kg for sediments and amphipods (Cai et al., 1996). The recovery from spiked pore water was 92% (n = 6), and the RSD for duplicate analyses was 24% (n = 6 pairs). For amphipods, the MeHg concentration in DORM was 356 ± 34 ng/L, close to the expected value of 355 ± 56 ng/L (100% recovery; n = 4). The average MeHg concentration in pore water samples spiked with 250 ng/L was 244 ± 7 ng/L (98% recovery; n = 4) and the RSD was 20% (n = 2 pairs). The recovery for spiked sediment samples was 108% (n = 15), with average RSDs of 6% for both duplicate spiked samples (n = 4 pairs) and duplicate unknowns (n = 2 pairs). 2.4. Data analysis Data were analyzed with GraphPad Prism 5.0 (GraphPad Software, San Diego, CA). Normality was assessed by a Shapiro–Wilk test, and homogeneity of variance by a Bartlett's test. Because data were skewed, log transformations were applied to attain normal distributions and homogenous variances, except for %MeHg in sediments. Differences among sites and zones in geometric mean Hg concentrations were assessed by analysis of variance with log-transformed data, and between specific sites by Tukey's Post Hoc test. An analysis of covariance (ANCOVA) described the effect of perch length on THg concentrations so that comparisons of THg among samples were based on estimated concentrations at a common total length of 150 mm. Correlations among variables were assessed with a Pearson correlation matrix and examples shown as scatter plots. Throughout the results and discussion, references to significant differences signify that p b 0.05. 3. Results 3.1. Sediments Ninety-six sediment cores were collected from 27 sites (Table SI-1) and the mean THg concentrations in the most contaminated surficial sediments (top 5 cm) ranged from 330 to 1528 μg/kg dw at sites close to Cornwall (Fig. 3A). Concentrations at north shore sites decreased eastward to less than 50 μg/kg dw, except for site 16, the furthest east (108 μg/kg dw). South shore sites showed a similar west-to-east decrease in sediment THg concentrations; at most sites, concentrations

Table 1 Characteristics (mean and 95% confidence limits) of yellow perch captured in three zones of Lake Saint Francis. Means sharing the same letter are not significantly different (p N 0.05). Zone

Cornwall

North

South

N Total Length (mm) Weight (g) Age (y) K1 Geometric mean total Hg (uncorrected for size) % N 260 μg/kg total Hg2 % N 330 μg/kg total Hg3

31 134 (123–145)a 30 (24–37)a 2.1 (1.7–2.5)a 1.11 (1.08–1.14)ab 115 (95–140)a 6.5 6.5

216 161 (157–165)b 51 (47–55)b 2.8 (2.6–3.0)b 1.11 (1.09–1.12)a 184 (173–195)b 21.4 11.2

220 153 (149–158)b 45 (42–49)b 2.5 (2.3–2.6)a 1.09 (1.08–1.11)b 79 (73–85)c 0.9 0.5

1

K = Condition Factor = (100 × Weight) / (total length)3. Ontario Fish Consumption Advisory Limit for sensitive populations (OME, 2011). 3 Threshold dose in fish muscle associated with sublethal toxicity to fish (calculated from an estimated threshold for whole-fish Hg Beckvar et al., 2005) using a conversion equation from Peterson et al. (2007). 2

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Sediment THg (ug/kg dw)

222

10000

A

1000 100 10 1

Sediment MeHg (ug/kg dw)

1

3

4

5

7

8

9 10 11 12 13 14 15 16 2 27 26 25 6 24 23 22 21 20 19 18 17

100

B

10 1 0.1

ND

ND

0.01

Pore water MeHg (ng/L) Amphipod MeHg (ug/kg dw)

1 1000

200

3

4

5

7

8

9 10 11 12 13 14 15 16 2 27 26 25 6 24 23 22 21 20 19 18 17

C 100 10


1 1

3

4

5

7

8

NDND

NDNDND ND

9 10 11 12 13 14 15 16 2 27 26 25 6 24 23 22 21 20 19 18 17

D 100

ND 0 1

3

4

5

7

8

9 10 11 12 13 14 15 16 2 27 26 25 6 24 23 22 21 20 19 18 17

300

Perch Hg (ug/kg ww)

ND

OMOE Consumption guidelines (260µg/kg)

E

200 100 0 1

3

4

Cornwall

5

7

8

9 10 11 12 13 14 15 16 2 27 26 25 6 24 23 22 21 20 19 18 17

North Shore

South Shore

Upstream-downstream

Upstream-downstream

Fig. 3. Geometric mean concentrations and 95% confidence limits of total Hg (THg) in sediments (A), methyl Hg (MeHg) in sediments (B), MeHg in pore water (C), MeHg in amphipods (D), and THg in yellow perch (E) from 27 sampling sites grouped by zone in L. St. Francis, St. Lawrence River. Perch Hg concentrations were not corrected for differences among sites in fish size. Sample sizes range from 1 to 5 for all variables except fish, where n = 10–22; where there are no error bars, n = 1. Sites with no bars indicate no data (ND) or concentrations less than the detection limit in pore water (bDL; 0.02 ng/L). Bars are filled to indicate low (white), medium (hatched), and high (black) values.

were less than 50 μg/kg dw, except at sites 2, 23, and 24, where concentrations ranged from 114 to 195 μg/kg dw. Geometric mean sediment THg concentrations at sites near Cornwall were 34-fold higher than concentrations averaged over all sites on the north or south shores (Fig. 4A); there was no difference between north and south shore geometric mean THg concentrations. The pattern of MeHg concentrations in sediments was similar to that of THg when considered site-by-site (Fig. 3A vs B), and the two measures were strongly correlated (Table 2). The highest concentrations were found at sites 1–5 near Cornwall (1.0–4.3 μg/kg dw), declining eastwards along the north (4.3–0.03 μg/kg dw) and south shores (1.01–0.04 μg/kg dw). Among zones, the pattern of MeHg distribution was identical to that of THg, with geometric mean concentrations at Cornwall significantly higher than at north and south shore sites, and no difference between north and south shore sites (Fig. 4A vs B). There were 63 pore water samples from 17 sites analyzed for MeHg, because insufficient pore water was extracted for analysis at sites 13, 14, 19, 20, 21, 22, and 24. Concentrations at sites 10, 12, 15 were below

the DL, so that concentrations of MeHg ranged from b0.02 ng/L to 69 ng/l (site 6). Where MeHg was measurable, there were no systematic trends among sites in concentration (Fig. 3C), but overall, geometric mean concentrations in pore water from south shore sites were 4-fold higher than at Cornwall (p b 0.05; Fig. 4C), a trend opposite to that for THg and MeHg; the geometric mean concentration across north shore sites was intermediate. Sediment THg and MeHg concentrations were both negatively correlated to pore water MeHg concentrations (Table 2), and there was no relationship between % MeHg in sediment and MeHg in pore water. Sediment organic matter followed a similar pattern of distribution as THg and MeHg (Figure SI-1), which were positively correlated to %LOI (Table 2; Fig. SI-2). The highest values for %LOI were measured at sites 1 to 5 near Cornwall, with the highest mean (15.1%) at site 3 (Fig. SI-1). Downstream of site 5, %LOI decreased and remained under 10% at north and south shores. Sediment grain size showed the reverse pattern, with the lowest mean sizes at sites 1 to 5 closest to Cornwall, ranging from 15.1 to 25.2 μm, and increasing from west to east on the north shore; there was no consistent trend in grain size on the south

P.V. Hodson et al. / Science of the Total Environment 494–495 (2014) 218–228

Sediment THg (ug/kg dw)

10000

a

A

1000

100

10

Sediment MeHg (ug/kg dw)

10.0

b 8

b

44

44

B

a

1.0

b 0.1

9

b

44

41

Pore water MeHg (ng/L)

100.0

10.0

a

a,b

b

C

were analyzed for THg only when sufficient material remained after MeHg analyses. Due to the low weight of amphipods, replicates at many sites were combined for a single sample, limiting statistical analyses to comparisons among zones. The highest concentrations of MeHg (Fig. 3D) and THg (Fig. SI 3A) in amphipods were found at sites along the north shore, particularly at site 7. The lowest proportion of THg concentrations that was MeHg (48%) was measured at site 3 (Cornwall; Fig. SI 3B), but overall there were no systematic trends. The concentrations of MeHg in amphipods were not correlated to sediment THg and MeHg, or to pore water THg and MeHg concentrations (Table 2). As sediment grain size increased, the concentration of THg in amphipods decreased but there was no relationship to %LOI. Conversely, as sediment THg concentrations and %LOI decreased, the %MeHg in amphipods increased, but there was no relationship between MeHg concentrations in amphipods and %LOI or grain size of the sediments. Concentrations of MeHg in amphipods showed patterns of distribution among zones that were different from those in sediments and sediment pore water. Concentrations appeared higher at north shore sites than at sites close to Cornwall or sites on the south shore (Fig. 3D), but the difference among zones was significant only for the north shore–south shore comparison (Fig. 4D). 3.3. Perch

1.0

0.1

Amphipod MeHg (ug/kg dw)

100

8

a,b

15

16

a

D b

6

24

19

10 200

Perch Hg (ug/kg ww)

223

b

E

a 150

c

100 50

31

215

220

Fig. 4. Geometric mean concentrations and 95% confidence limits of total Hg (THg) in sediments (A), methyl Hg (MeHg) in sediments (B), MeHg in pore water (C), MeHg in amphipods (D), and size-corrected THg in 150 mm yellow perch (E) in three zones of Lake Saint Francis, Saint Lawrence River. The numbers within bars are sample sizes; treatments sharing the same letter are not statistically different (p N 0.05).

shore. Sediment THg and MeHg concentrations were negatively correlated to grain size. The MeHg concentrations in pore water were also positively correlated to %LOI, but not to grain size. 3.2. Amphipods Twenty-two amphipod samples were analyzed for THg and 49 for MeHg; MeHg is the form of mercury that bioaccumulates, and samples

A total of 467 yellow perch were sampled and analyzed for THg. Across all sites, there were significant differences in fish age, total length, weight, and condition factor (K = [100 × weight (g)] / [total length3 (mm)]), which ranged from 1 to 6 y, 92 to 246 mm, 9 to 176 g, and 0.72 to 1.42, respectively (Table SI-1). Perch from sites 1 (Cornwall) and 18 (south shore) were notably younger and smaller than those from other sites, but within zones, there were no systematic trends in fish characteristics (Table SI-1). Among zones, perch from north shore sites were larger on average than those at Cornwall or the south shore (Table 1), and although the average K of fish at Cornwall was statistically greater than K of fish from the north and south shores, the differences were small (b 5%). The weight–length relationships of perch from all zones followed a single distribution (Fig. SI-4), as did size–age distributions (Fig. SI-5), indicating no differences in growth rates among zones; the differences in size and K were likely driven by younger fish at Cornwall (Fig. SI-5). Concentrations of THg in perch were strongly correlated to total length at all three zones, (Fig. 5), as well as to age and to other measures of size (standard and fork lengths, weight; data not shown). There were no significant differences in slopes of THg–length relationships among zones, as indicated by ANCOVA; i.e., there was a common rate of increase of tissue THg with increasing fish size. Nevertheless, the data demonstrate considerable variance, likely due to significant site-to-site differences in THg (Fig. 3E). Among sites, the pattern of THg concentrations in perch was similar to that of amphipods, and quite different to that of sediments. Perch from sites along the north shore contained higher concentrations of THg than fish from Cornwall and from the south shore (Fig. 3E). The range of concentrations across all 467 fish was 32 to 498 μg/kg ww, but the highest geometric mean THg concentrations were found at sites 4, 7, 8 and 12, ranging from 207 to 236 μg/kg ww respectively; mean concentrations at south shore sites ranged from 42 to 109 μg/kg ww (Fig. 3E). Among zones, geometric mean concentrations in north shore perch were 50% higher than in Cornwall perch and more than 2-fold higher than in south shore perch (p b 0.05; Table 1). When the comparisons among zones were standardized by ANCOVA to a common total fish length of 150 mm, the pattern of Hg distribution and differences among zones remained the same (Fig. 4E). Yellow perch THg concentrations were not correlated to sediment THg (Fig. 6A) and MeHg concentrations, %MeHg in sediments, MeHg in pore water, %LOI, or sediment grain size (Table 2). The only significant correlations were between perch THg and amphipod THg and MeHg

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Table 2 Correlation matrix based on mean values for each site.

log MeHg Sediment % MeHg Sediment log grain Size log %LOI

log MeHg Pore water logTHg Amphipod log MeHg Amphipod % MeHg Amphipod log perch THg

r p n r p n r p n r p n r p n r p n r p n r p n r p n

log THg sediment

log MeHg sediment

%MeHg sediment

log grain size

log %LOI

log MeHg pore water

log THg amphipod

log MeHg amphipod

% MeHg amphipod

0.928 b0.001 25 −0.566 0.003 25 −0.668 b0.001 27 0.852 b0.001 26 −0.617 0.008 17 0.299 0.243 17 −0.088 0.668 26 −0.678 0.004 17 0.044 0.829 27

−0.288 0.163 25 −0.646 b0.001 25 0.897 b0.001 25 −0.719 0.001 17 0.349 0.185 16 −0.068 0.752 25 −0.704 0.003 16 0.038 0.856 25

0.308 0.134 25 −0.317 0.123 25 0.227 0.380 17 0.129 0.635 16 0.246 0.247 25 0.420 0.119 16 0.141 0.500 25

−0.729 b0.001 26 0.469 0.057 17 −0.542 0.025 17 −0.066 0.749 26 0.460 0.073 17 0.057 0.777 27

−0.843 b0.001 17 0.433 0.082 17 0.053 0.800 25 −0.624 0.010 17 0.036 0.860 26

−0.226 0.531 10 −0.137 0.599 17 0.204 0.599 10 −0.057 0.827 17

−0.309 0.245 17 0.561 0.019 17

0.196 0.468 17 0.751 b0.001 26

0.286 0.283 17

Hg = mercury; THg = total mercury; MeHg = methyl mercury; %LOI = percent loss on ignition. Bold text signifies a statistically significant correlation (p b 0.05).

concentrations; the latter correlation was strongest (Fig. 6B). Normalization of THg and MeHg in sediments to organic carbon (i.e., concentration/(%LOI/100)) did not improve correlations to Hg concentrations in amphipods and perch (data not shown). Assuming that %LOI of sediments predicts concentrations of organic matter in pore water, pore water MeHg concentrations were also normalized, but again with no improvement in correlations to Hg in biota. The frequency with which concentrations of THg in perch exceeded the OME's fish consumption guideline of 260 μg/kg ww for sensitive populations also varied with zone (Table 1). The probability of catching a perch in excess of the guideline was 21% along the north shore, three times the probability at Cornwall, and 25 times that probability on the south shore; at individual sites, the probability was as high as 45% (Fig. SI-6). 4. Discussion There was a good correspondence between THg concentrations in surficial sediments measured in 2008 by Pelletier, (2010) and by the present study (Fig. SI-7). Concentrations were not distributed evenly throughout LSF, averaging 30 to 35 times higher at sites closest to Cornwall compared to most sites along the north or shore shores. The distribution of different forms of sediment mercury was also not consistent. Sediment MeHg followed the same pattern as THg, but differences among zones were much smaller, with concentrations at Cornwall only 6 to 8-fold higher than in the other two zones. In contrast, the distribution of pore water MeHg was distinctly different, with concentrations 3- and 4-fold less at Cornwall than at north and south shore sites, and highest where THg concentrations were lowest. The pattern of Hg concentrations in biota was markedly different to that in sediments. Amphipods and fish were most contaminated at north shore sites; Hg concentrations at Cornwall and south shore sites were 25% and 43% lower, respectively. The differences among zones in patterns of Hg contamination were also reflected in the absence of statistical correlations between measures of Hg in sediment and in biota. While THg and MeHg concentrations in sediment were positively

correlated, they were both negatively correlated to sediment pore water MeHg. None of these measures were correlated to concentrations of MeHg in amphipods or THg in perch. The apparent disconnect between Hg concentrations in biota and sediments challenges traditional concepts of bacterial-mediated transfer of sediment Hg to aquatic food webs. The expectation was that microbial methylation of sediment Hg (Merritt and Amirbahman, 2009; Avramescu et al., 2011), would increase concentrations of MeHg in pore water, and then in benthos and fish via biomagnification through one or more trophic transfers (Watras et al., 1998; Lawrence and Mason, 2001; Skyllberg et al., 2007). Because MeHg is the form considered most bioavailable, correlations between sediment and biota Hg concentrations should be strongest for measures of MeHg; however, no correlations were evident. The absence of statistical links suggests that the bioavailability of legacy Hg in sediments is very low, which accords with the observation that the extent of Hg binding and complexation increases with sediment aging; older deposits of sediment Hg are less bioavailable than freshly deposited Hg (Munthe et al., 2007). Thus, the extent of legacy sediment contamination by Hg was unrelated to current concentrations in fish, and the spatial distribution of Hg in sediments was not a reliable indicator of the spatial distribution of risks to fish consumers. The absence of statistical links between sediment and biota Hg concentrations might be due to the small sample size at Cornwall (N = 2 sites). However, the results for sediment and biota Hg corresponded well with those of previous studies at these same sites (Fowlie et al., 2008; Razavi et al., 2013), and there was no correlation between sediment and fish Hg concentrations when all 27 sites were considered. It is unlikely, therefore, that the small number of sites at Cornwall affected the overall conclusions of the present study. While there is little evidence for Hg transfer from sediments to biota, there was evidence for trophic transfer, as indicated by strong correlations between MeHg in amphipods and THg in perch. In LSF, amphipods are a dominant component of juvenile perch diets (Yanch, 2007), and higher Hg concentrations in fish relative to amphipods (2.5-, 2.5-, and 2.0-fold higher at Cornwall, north shore and south shore sites,

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3

log total Hg (ng/g ww)

North Shore

2.5

2 South Shore

Cornwall

1.5

1 50

100

150

200

250

Total length (mm) Fig. 5. The variation of log total Hg concentration with total length of yellow perch from three zones of Lake Saint Francis, Saint Lawrence river. An analysis of covariance demonstrated that slopes were parallel and described by the equation: log THg = 1.4524 + 0.0042 × total length (mm) + zone constant. The zone constants (and sample sizes) were: Cornwall: 0.0507 (31); North shore: 0.1442 (214); South shore: −0.1948 (221); the R2 was 0.59, and the regression was significant (p b 0.001). The vertical line indicates the log mean THg concentration of perch at a common length of 150 mm.

respectively) were consistent with biomagnification (Rolfhus et al., 2011). The pattern of Hg concentrations in amphipods provides a much better basis for predicting spatial and temporal variations in Hg concentrations of perch than measures of sediment Hg.

log perch THg (µg/kg)

3.2

A

3.0 2.8 2.6 2.4 2.2 2.0 0

1

2

3

4

log sediment THg (µg/kg)

log perch THg (µg/kg)

3.5

B

3.0

2.5

2.0 1.0

1.5

2.0

2.5

log amphipod MeHg (µg/kg) Fig. 6. Log total mercury (THg) concentrations in yellow perch versus (A) log THg in sediments (n = 32, r = 0.04; p = 0.83) and (B) log methylmercury (MeHg) in amphipods (n = 26, r = 0.75, p b 0.0001).

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One reason for a low bioavailability of sediment Hg may be binding or complexation of inorganic Hg to sediment organic matter and minerals (Kim et al., 2008; Canário et al., 2008), which reduces the transfer of Hg to methylating bacteria. Similarly, while MeHg is hydrophobic and highly bioavailable, Hg binding to humic compounds with hydrophobic character would reduce uptake by benthos (Lawrence and Mason, 2001; Hammerschmidt and Fitzgerald, 2004; Canário et al., 2008). In LSF, sediment THg and sediment organic matter were strongly correlated, and concentrations of organic matter were greatest at contaminated sites near Cornwall. There was also a strong negative correlation between sediment THg concentrations and grain size, likely because fine clay particles bind inorganic Hg and settle in areas of low current velocity (Rudd et al., 1983; Lepage et al., 2000). Hence, perch THg concentrations should be highest where %LOI is lowest and grain size is highest, i.e., in areas with the least potential for Hg complexation. Sizeadjusted THg concentrations in fish at Cornwall were lower than at north shore sites, consistent with less bioavailable Hg. Nevertheless, there were no correlations between %LOI and sediment grain size and THg in fish. Normalizing sediment Hg concentrations to organic matter concentrations did not improve correlations to Hg in biota (data not shown), suggesting that other factors control Hg concentrations in fish. For example, differences in the quality of organic matter, especially the presence of organic sulfur components such as thiols, can affect the availability of Hg to methylating bacteria (Rydberg et al., 2008; Shchukarev et al., 2008; Schartup et al., 2014). Canário et al. (2008) found a positive correlation between THg and organic S in LSF sediment solids, suggesting that variations of THg concentrations in these sediments can be strongly influenced by organic sulfur compounds at the sediment–water interface. Apart from sediment characteristics, the variations of Hg concentrations in fish among sites could be related to water quality and hydrologic variables unique to each site. Among 131 Adirondack and Catskill lakes in New York, there was significant variability in Hg content among four species of fish, including yellow perch (Simonin et al., 2008). The source of Hg to all lakes was the same, i.e., atmospheric transport. However, fish of the same species from adjacent lakes had markedly different Hg concentrations that were correlated to water quality characteristics, control of water levels by dams, and proximity to wetlands. Alternative explanations for the spatial distribution of Hg in perch from LSF include variations among sites in perch demographics that affect bioaccumulation (Munthe et al., 2007), and additional sources of Hg, including the atmosphere. Perch were sampled over a large geographic area, from habitats that differed in sediment characteristics, proximity to tributaries, vegetative cover, and current strength. Despite upstream–downstream trends of decreasing Hg concentrations in perch, differences between adjacent sites separated by 1 to 8 km were often more than 2-fold. The extent of spatial differences were similar to those previously observed in perch and spottail shiners at contaminated sites in Cornwall (Choy et al., 2008; Fowlie et al., 2008). In May, 2004–05, the variation in Hg concentrations among perch at site 1 was also enhanced by the brief appearance of mature spawning perch with much lower Hg concentrations than locally-resident immature perch (Fowlie et al. 2008). In the present study, perch were sampled from July to September to avoid migration bias associated with spawning. Mercury concentrations in fish increase with size, reflecting the progressive accumulation of Hg with age and, for some species, a transition in trophic status as growing fish seek larger prey. Perch sampled at Cornwall in 2004–05 demonstrated similar size–age relationships (i.e., growth rates) among sites, a progressive increase in Hg concentrations with age, and significant differences among sites in geometric mean Hg concentrations when total length was standardized to 148 mm (Fowlie et al., 2008). A truncated age frequency distribution suggested that yellow perch abandoned site 1 at Cornwall at age 3 +, likely because of food limitations when they switched diets from invertebrates to small

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fish. The characteristics of perch in the present study were virtually the same. Although fish from sites 1 and 18 were significantly smaller than average, there were no significant differences in weight-length and size–age relationships among sites, indicating that the truncated age distribution at site 1 did not affect condition or growth rates. Thus, demographic characteristics did not vary sufficiently among sites to explain the greater accumulation of Hg by perch at north shore sites compared to Cornwall or the south shore. The significant differences across all sites in perch THg concentrations, as high as 5.5-fold, and the lower concentration of Hg in south shore perch may be related to the many tributaries to LSF between Cornwall and Montreal (Fig. 2). Concentrations of THg and MeHg in tributaries ranged from 0.69 ng/L–1.62 ng/L and 0.13 ng/L–0.25 ng/L, respectively, several-fold greater than average concentrations (0.47 ng/L THg and 0.03 ng/L MeHg) in the main stem SLR (mainly L. Ontario water; Ridal et al., 2010). The size and topography of watersheds, ratios of watershed-to-surface water area, land cover, and land use (forest vs agriculture) are important determinants of Hg flux to surface waters (Munthe et al., 2007). On the north shore, Hg concentrations in perch were higher than average at sites 7, 8, 12, and 13, adjacent to Fraser Creek, the Raisin River, Wood Creek, and Rivière Beaudette, respectively. North shore tributaries drain agricultural and reclaimed wetlands that flood seasonally and that are dried by tile drainage, a cycle that enhances microbial production of MeHg (Munthe et al., 2007; Rolfhus et al., 2011). For example, headwater wetlands, particularly bogs, are an important source of THg, MeHg, and DOC to the Raisin River (Maharaj, 2008). Bogs contain peat, which accumulates Hg from long-term atmospheric deposition (Shotyk et al., 1992). When wetlands are disturbed, e.g., when water levels are altered, Hg can be released from peat by drying and rewetting. Drying promotes oxidation of sulfur compounds, the production of sulfuric acid, lower porewater pH, and mobilization of metals with rewetting (Tipping et al., 2003). High loadings of nutrients and suspended sediments from eroding soils enhance the rate of Hg methylation (Munthe et al., 2007; Skyllberg et al., 2007; Simonin et al., 2008) and the net flux of MeHg from sediments to food webs. In contrast, south shore tributaries drain mixed use watersheds that include the forested slopes of the Adirondack Mountains and proportionately less wetland. South shore tributaries contain markedly higher Hg and DOC and lower dissolved mineral and nutrient concentrations than in north shore tributaries (Ridal and Twiss, 2011). High DOC concentrations reduce Hg bioavailability (Rolfhus et al., 2011; Lawrence and Mason, 2001), although they may also increase Hg methylation (Hsu-Kim et al., 2013). Additional research is needed to compare total loadings and the forms and bioavailability of Hg between north and south shore tributaries to elucidate the role of tributary inputs in determining Hg concentrations in fish from the nearshore areas of the SLR. Sediment re-suspension also increases the release of MeHg from sediment pore waters (Kim et al., 2008). Porewater MeHg concentrations were higher than in surface waters, ranging from b 0.02 to 23.8 ng/L (geometric mean 10.7 ng/L, n = 15) at north shores sites. Similar values at south shore sites (geometric mean 13.3 ng/L, n = 16) were comparable to other recent porewater measurements of MeHg in Saint Lawrence River sediments (Canário et al., 2008; Fathi et al., 2013). Nevertheless, onshore winds, wave action, and sediment re-suspension are highest on the south shore (Lepage et al., 2000), inconsistent with the lowest concentrations of Hg in perch and amphipods. Another potential source of mercury to LSF is atmospheric deposition. Gaseous elemental mercury (GEM) is efficiently transported around the globe. GEM enters aquatic systems when small amounts are transformed to reactive Hg(II) and scavenged in rainfall and particulate dryfall (Morel et al., 1998). Sources of atmospheric Hg deposition may be local or regional, and may have significant impacts on concentrations in local water bodies and biota, especially in pristine lakes (Wiener et al., 2006). Poissant et al. (2005) reported atmospheric

deposition of Hg to LSF from samples collected at St. Anicet, Qc, on the south-east shore of LSF. Rates of deposition and concentrations in rain and particulates were consistent with known regional (Great Lakes) and local sources. For example, the highest concentrations were associated with air masses from the NE, reflecting the impact of the major urban centre of Montreal, Qc. Air masses from the NE occurred only 17% of the time during the annual monitoring, while flow from the SW occurred 50% of the time and represented the lowest atmospheric Hg concentrations. Given the scale of sediment and biota sampling in the present study (sites approximately 1– 8 km apart), the pattern of downstream distribution of sediment Hg concentrations from industrial sources at Cornwall, and the clear north–south differences in biota Hg concentrations, it is unlikely that local or regional atmospheric deposition accounted for the differences in Hg concentrations in yellow perch. The listing of LSF as an AOC responded to concerns for potential risks to humans and wildlife consumers of Hg-contaminated fish. High concentrations in perch along the north shore might still be caused by an active source of Hg at Cornwall. In 2008, ‘fugitive’ Hg emissions from old industrial sewers were discovered and controlled (Ridal et al., 2010), which could explain the 40% decrease in size-corrected Hg concentrations in site 1 perch between 2004 and 2005 (220 μg/kg ww; Fowlie et al., 2008) and 2008 (133 μg/kg ww; present study). In addition, the rates of Hg methylation and release from Cornwall sediments were not constant, because concentrations of pore water MeHg declined to non-detectable from spring to late summer (Razavi et al., 2013). Net rates of Hg methylation in surface sediments (methylation minus demethylation) are low at Cornwall and a negligible diffusion of MeHg from sediments into overlying waters (Fathi et al., 2013). Fresh inputs of Hg are required to increase net methylation (Avramescu et al., 2011), consistent with a decline in Hg efflux as sediments age (Munthe et al., 2007). Thus, Hg contamination of fish likely reflects fresh sources of Hg to LSF from tributaries and from the atmosphere, and not transfer of legacy Hg from sediments. This finding is similar to that of two different studies of marine coastal ecosystems. Across a gradient of Hg contamination in 10 coastal estuaries on the NE coast of the US, concentrations of MeHg in two species of forage fish were most closely related to concentrations in suspended particulates, and not to sediment concentrations (Chen et al., 2014). Similarly, a modeling study of Hg transfer and mass balance for Passamaquoddy Bay, Canada, found no strong links between Hg in sediments and biota (Sunderland et al., 2010). In both cases, the most important source of Hg to fish appeared to be Hg in the water column. The data presented in these studies, and in the present study, do not support programs of active sediment remediation to reduce Hg concentrations in fish, and spatial or temporal trends in sediment concentrations do not provide insight on the spatial distribution of risks to consumers of fish. Nevertheless, existing controls on shoreline development to limit the disturbance of sediments at contaminated sites at Cornwall may be justified to reduce the potential for remobilization of sediment Hg. Monitoring of Hg in fish is essential for understanding the current risks to consumers of fish. The results of the present study, and longterm declines of Hg concentrations in fish throughout the SLR (Neff et al., 2013), suggest that fish are exposed to regional as well as local sources of Hg. The best current indicator of risk is analysis of fish tissues, but the high degree of spatial variance of Hg in yellow perch is an obstacle to accurate risk assessment. Risks appear least on the south shore because only one fish in a hundred, across all sizes, exceeded the guideline for fish consumption of 260 μg/kg ww (OME, 2011). In contrast, on the north shore, there was a 1 in 5 chance of catching a perch that exceeded the guideline, and 11% of perch exceeded the estimated threshold muscle concentration of 330 μg/kg ww associated with sublethal toxicity to fish (Beckvar et al., 2005; Table 1). At Cornwall, the site of greatest sediment contamination, the risk of catching a contaminated fish was only 1 in 15, in part because the fish were younger and smaller than at other sites. Hence, monitoring of fish at only one site within the LSF river block

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could misrepresent risk at other sites by up to 25-fold. Single samples create a strong potential to under-protect fish consumers at some sites, and over-protect at others, with impacts on sports and commercial fisheries, tourism, and other forms of river-based development. We strongly support a recommendation by Neff et al. (2013) for more intensive sampling to represent the spatial distribution of Hg risks in large lacustrine lakes with complex hydrology and numerous potential sources. Appropriate strategies might be to first describe the variations among potential sampling sites (or sites where fishing is common), and then sample either a smaller, but representative array of sites, or simply the most contaminated site. The first alternative would provide consumers an idea of relative risk among sites, while the latter would be less expensive but provide very conservative ‘worst case’ estimates of risk. Either approach would be site specific, reflecting ecosystem characteristics, the intensity and location of fishing by consumers, and resources available for monitoring. Important design factors would include site selection and capture of a sufficient fish to normalize Hg concentrations to a common size for comparisons among sites and years. While more intensive sampling would increase the cost of surveys, the relatively slow rates of change of Hg in fish would support less frequent sampling.

5. Conclusions In summary, there was little connection between mercury concentrations in highly contaminated surficial sediments of Lake St. Francis and concentrations of mercury in aquatic biota, which challenges management models that connect past industrial contamination by mercury with current risks to consumers of mercury-contaminated fish. Contamination of fish is likely related to current loadings of mercury from tributaries or the atmosphere. Management to reduce the risk to human consumers of fish should focus on controlling current loadings and not on remediating sediments contaminated with ‘legacy’ mercury. Programs to monitor mercury in fish should explicitly measure spatial variations in the distribution of mercury-contaminated fish to avoid over- or under-protecting consumers of fish by sampling only at one site.

Conflict of interest The authors declare that they have no potential sources of conflict of interest that would influence their objectivity.

Acknowledgments Technical Support was supplied by D. Flood, S. Eveno, A. LeNevaneen, C. Mizon, M. Derouin, D. Ridal, D. Filion, L. St. Marseille, J. Szwec, R. Gauthier, C. Veenstra, A. Gulka, L. Cowell, and L. St. Pierre. Dr. J. Marty provided advice about sampling designs and assisted with statistical analyses, and Dr. E. Yumvihoze analyzed MeHg in sediments at the University of Ottawa mercury analysis laboratory, directed by Prof. A. Poulain. Analyses of %LOI and grain size were provided by the Analytical Services Unit and Dr. Scott Lamoreux, respectively, of Queen's University. This research was funded by a Best-in-Science research contract to JJR and PVH from the Ontario Ministry of the Environment (78550). Financial support to KN was provided by Queen's University and to MB by the Student Work Experience Program of the Government of Canada.

Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2014.06.137.

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