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Sources and Fate of Organochlorine Pesticides in North America and the Arctic

by

Liisa M. Jantunen

A thesis submitted in conformity with the requirements

for the degrees of Doctor of Philosophy

Graduate Department of Department of Chemical Engineering and Applied Chemistry

University of Toronto

©Copyright by Liisa M. Jantunen 2010

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Sources and Fate of Organochlorine Pesticides in North America and the Arctic

Liisa M. Jantunen

Doctor of Philosophy

Department of Chemical Engineering and Applied Chemistry University of Toronto

2010

ABSTRACT

Atmospheric transport and air-water exchange of organochlorine pesticides (OCPs) were investigated in

temperate North America and the Arctic. OCPs studied were hexachlorocyclohexanes (HCHs, α-, β- and γ-isomers),

components of technical chlordane (trans- and cis-chlordane, trans-nonachlor), dieldrin, heptachlor exo-epoxide and

toxaphene. Air and water samples were taken on cruises in the Great Lakes and Arctic to determine concentrations

and gas exchange flux direction and magnitude. The Henry’s law constant, which describes the equilibrium

distribution of a chemical between air and water, was determined for several OCPs as a function of temperature and

used to assess the net direction of air-water exchange. Air samples were collected in Alabama to investigate southern

U.S. sources of OCPs. Chemical markers (isomers, and enantiomers of chiral OCPs) were employed to infer sources

and trace gas exchange. Elevated air concentrations of toxaphene and chlordanes were found in Alabama relative to

the Great Lakes, indicating a southern U.S. source. Profiles of toxaphene compounds in air were similar to those in

soil by being depleted in easily degraded species, suggesting that soil emissions control air concentrations. Gas

exchange fluxes in the Great Lakes indicated near-equilibrium between air and water with excursions to net

volatilization or deposition. Net volatilization of α-HCH from the Arctic Ocean was traced by evasion of non-racemic

α-HCH into the atmosphere.

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ACKNOWLEDGMENTS

This thesis would not have been possible without the support, guidance and assistance of family and colleagues.

I would like to thank Terry Bidleman, for being my supervisor and supporting me over the years and years that it

took this thesis to be completed.

Tim MacNaughtan, my husband, for telling me I had to finish.

I would like to thank Paul Helm and Andi Leone for their support in the laboratory.

I would also like to thank all the people who made the field studies possible. BERPAC-93: Alla Tsyban, Jackie

Grebmeier, Cliff Rice and crew and fellow scientists from the R/V OKEAH. AOS-94: Rob Macdonald and the

crew and fellow scientists from the CCGS Louis S.St. Laurent. TNW-99: Henrik Kylin and Swedish Polar

Secretariat for ship time and the crew and fellow scientists of the Louis S. St. Laurent. Great Lakes cruises: Janine

Wideman, Paul Helm and Jeff Ridal for help during sampling and the crew and fellow scientists of the CCGS

Limnos.

I would also like to thank Environment Canada and the Northern Contaminants Program for financial support.

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ABSTRACT ……………………………………………………………………………………… ii

ACKNOWLEDGEMENTS ……………………………………………………………………… iii TABLE OF CONTENTS ………………………………………………………………………… iv LIST OF TABLES .………………………………………………………………………………. vi LIST OF FIGURES ……………………………………………………………………………… viii Chapter 1.

1.1 CONCLUSIONS ……………………………………………………………………. 1 1.2 RECOMMENDATIONS …………………………………………………………… 3

Chapter 2: INTRODUCTION ………………………………………………………………… 4 Chapter 3: COMPOUNDS INVESTIGATED 3.1. Toxaphene ………………………………………………………………………….. 5 3.1.1 Toxaphene Congeners ……………………………………………………. 6

3.2. Cyclodienes …………………………………………………………………………. 8 3.2.1 Chlordane …………………………………………………………………. 8 3.2.2 Dieldrin …………………………………………………………………… 9 3.3. Hexachlorocyclohexanes …………………………………………………………… 10 Chapter 4: RELEVANT PROPERTIES AND PROCESSES 4.1. Physicochemical Properties ………………………………………………………… 11 4.1.1 Henry’s Law Constants …………………………………………………… 11 4.1.2 Determination of HLCs by the Gas Stripping Method …………………… 13 4.2. Air-Water Gas Exchange 4.2.1 Fugacity and the Net Exchange Direction ………………………………… 15 4.2.2 Uncertainty in Fugacity Ratios ……………………………………………. 15 4.2.3 Rate of Gas Exchange: The Modified Two Film Model …………………. 15 4.3. Chemical Tracers of Exchange Processes 4.3.1. Isomer and Parent-Metabolite Pairs …………………………………….. 17 4.3.2. Chiral Compounds ………………………………………………………. 17 4.3.3. α-HCH Enantiomers …………………………………………………….. 18 4.3.4. Cyclodiene Enantiomers ………………………………………………… 20 4.3.5. Enantiomers in Source Identification and Exchange Processes ………… 21

Chapter 5: MATERIALS AND METHODS 5.1. Sampling Locations ………………………………………………………………… 22 5.2. Air Sampling ………………………………………………………………………... 22 5.3. Water Sampling……………………………………………………………………... 24 5.4. Samples Extraction and Cleanup/Fractionation ……………………………………. 24 5.5. Analysis Methods, Quantitative and Chiral ………………………………………... 24 5.6.Quality Control ……………………………………………………………………… 25 Chapter 6: SOURCES, TRANSPORT AND ENVIRONMENTAL OCCURRENCE 6.1. Southern Sources and Transport 6.1.1. Toxaphene ………………………………………………………………... 25

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6.1.2. Cyclodienes ………………………………………………………………. 37 6.1.3. HCHs ……………………………………………………………………... 29 6.2. Great Lakes 6.2.1. Air ………………………………………………………………………… 29 6.2.2. Water ……………………………………………………………………… 31 6.2.3. Air-Water Gas Exchange …………………………………………………. 32 6.2.4. Chiral Tracers of Gas Exchange ………………………………………….. 37 6.3. Arctic

6.3.1. Air …………………………………………………………………………. 37 6.3.2. Water ……………………………………………………………………… 39 6.3.3. Air-Water Gas Exchange …………………………………………………. 41 REFERENCES …………………………………………………………………………………… 43

Original Papers

1 Jantunen, L.M., Bidleman, T.F. 2006. Henry’s Law constants for hexachlorobenzene, p,p’-DDE and components of technical chlordane and estimates of gas exchange for Lake Ontario. Chemosphere 62, 1689-1696. ……………………………………………….. 61

2 Jantunen, L.M., Bidleman, T.F., Harner, T. 2000. Toxaphene, chlordane

and other organochlorine pesticides in Alabama air. Environmental Science and Technology 34, 5097-5105. ………………………………….………………………… 73

3 Jantunen, L.M., Helm, P.A., Bidleman, T.F. 2008. Air-water gas exchange of

chiral and achiral pesticides in the Great Lakes. Atmospheric Environment 42, 8533-8542 ………………………………………………………………………………... 91

4 Jantunen, L.M., Bidleman, T.F. 2003. Air-water gas exchange of toxaphene

in Lake Superior. Environmental Toxicology and Chemistry 22, 1229-1237……………. 117 5 Jantunen, L.M., Bidleman, T.F. 1996. Air-water gas exchange of hexachlorocyclohexanes

(HCHs) and the enantiomers of α-HCH in Arctic regions. Journal of Geophysical Research 101, 28837-28846, corrections ibid. 1997, 102, 19279-19282…………………… 135

6 Jantunen, L.M., Helm, P.A., Bidleman, T.F., Kylin, H. 2008. Hexachlorocyclohexanes

(HCHs) in the Canadian archipelago, 2. Air-water gas exchange of α, and γ-HCHs. Environmental Science and Technology 42, 465-470 and Supporting Information. ……… 158

7 Jantunen, L.M., Bidleman, T.F. 1998. Organochlorine pesticides and enantiomers

of chiral pesticides in Arctic Ocean water. Archives of Environmental Contamination and Toxicology 35, 218-228………………………………………………. 177

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List of Tables

Chapters 3-6 Table 1 Toxaphene nomenclature ………………………………………………………… 7 Table 2 Physical-chemical data for OCPs ………………………………………………… 12 Table 3 Distribution of OCPs EFs in background soils, % of total samples ……………. 21 Table 4 Atmospheric concentrations of OCPs in the Great Lakes and Arctic ……………. 35 Table 5 Water concentrations of OCPs in the Great Lakes and Arctic …………………… 36 B. Original Papers Paper I Table I-1. Parameters of log H=m/T +b and enthalpies of water-air transfer (∆Hwa) ………… 77

Table I-2. Comparison of Henry’s Law constants, Pa m3 mol-1 ………………………………. 78

Table I-3. Air (Ca, pg m-3) and water (Cw, pg L-1) concentrations from Lake Ontario, July 1998 used to calculate fugacity ratios. ……………………………………….. 79

Paper II Table II-1. OCs in Alabama air, January to October 1996 and May 1997, pg m-3 ± SD. ……… 87

Table II-2. Ratios of chlordane compounds in air, soil and technical chlordane. ……………… 92

Table II-3. Mean atmospheric concentrations (pg m-3) of OCs in the southern U.S. and Great Lakes regions. ………………………………………………………….. 95

Table II-4. Regression parameters of log P/Pa versus 1/T plots. ………………………………. 97 Paper III Table III-1. Concentrations of gas phase organochlorine pesticides in air, pg m-3. ………….. 108

Table III-2. Concentrations of dissolved organochlorine pesticides in surface water, pg L-1. ……………………………………………………………………………. 110

Table III-3. Comparison to other Great Lakes air measurements (pg m-3). ………………….. 113

Table III-4. Fugacity and flux calculations for the Great Lakes. ……………………………. 120

Table III-5. Enantiomer fractions of chiral organochlorine pesticides in surface water and air. 122 Paper IV Table IV-1. Water concentrations of dissolved toxaphene in Lake Superior, Great Lakes,

pg L-1 ……………………………………………………………………………… 135

Table IV-2. Atmospheric concentrations of toxaphene over the Great Lakes, August 1996 and May 1997, pg m-3, see Figure 1 for sample locations. ………………………… 136

Table IV-3. Fugacity ratio and flux calculations for cruises and annual predictions, see Figure 1 for sample locations. …………………………………………………… 140

Paper V Table V-1. Hydrographic information, concentration of α- and γ-HCHs and enantiomeric

Ratio (ER) of α-HCH in water ……………………………………………………. 157

Table V-2. HCH concentrations in surface water in sub-arctic and arctic regions …………….. 159

Table V-3. Concentrations of HCHs and enantiomeric ratio (ERs) of α-HCH in air ………….. 160 Table V-4. Fugacity ratio and flux calculations ………………………………………………. 162

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Paper VI Table VI-1. Atmospheric concentrations of HCHs during TNW-99 and at Resolute

Bay (RB)a, pg m-3 ………………………………………………………………. 175

Table VI-2. Air and water concentrations of α- and γ-HCH on TNW-99 …………………… 176

Table VI.3. Atmospheric concentrations and fluxes of α- and γ-HCH at Resolute Bay ……. 179

Table VI-4. Water concentrations of HCHs by zone, fugacity ratios and net fluxes ............... 180

Table VI-5. Fugacity and flux calculations for TNW-99 …………………………………….. 181

Paper VII Table VII-1. Hydrographic information and concentrations (pg L-1) of dissolved pesticides …… 195

in surface water.

Table VII-2. Average regional concentrations of pesticides in surface water (pg L-1) ………….. 201

Table VII-3. Enantiomeric ratios of chiral pesticides in surface water …………………………. 206

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List of Figures Chapters 3-6 Figure 1 Structures of toxaphene congeners ……………………………………………….. 5 Figure 2 Structures of cyclodiene OCPs …………………………………………………… 9 Figure 3 Structure of HCH ………………………………………………………………… 10 Figure 4 Triangular relationship between KOA, KOW and KAW …………………………… 11 Figure 5 Henry’s Law apparatus …………………………………………………………… 14 Figure 6 Enantiomers of α-HCH …………………………………………………………... 18 Figure 7 Enantiomers of trans-chlordane ………………………………………………… 20 Figure 8 Sample collection sites, AOS-94, Great Lakes 1996-2000 and TNW-99 ……… 23 Figure 9 Toxaphene chromatograms, showing air and soil from Alabama compared

to a standard …………………………………………………………………….. 28 Figure 10 α-HCH in air: EF versus air concentration for Lake Superior, August

1996 and May 1997 ……………………………………………………………… 34 Figure 11 EF with depth in the Arctic Ocean at four stations ……………………………… 42 Figure 12 EF of α-HCH in air from AOS-94 ………………………………………………. 43 B. Original Papers Paper I Figure I-1. Bubble stripping experiment at 10oC, for trans-chlordane (--▲--), ……………… 75

cis-chlordane (--■--) and trans-nonachlor(�●�). Figure I-2. Plots of Eq. (5) for HCB (a), trans-chlordane (b), trans-nonachlor (c) …………… 76

and p,p’-DDE (d). Figure I-3. Fugacity ratios for trans-chlordane, calculated with the HLCs determined in

this study, in comparison to literature values. The error bars are derived from propagation of errors (as in Sahsuvar et al., 2003) for this study only. ……… 79

Paper II Figure II-1 Concentration in pg m-3 (a) γ-HCH, (b) heptachlor, (c) trans-chlordane (TC) and trans-nonachlor (TN), (d) cis-chlordane (CC) and heptachlor exo-epoxide (HEPX), (e) dieldrin and oxychlordane (OXY), (f) p,p’-DDE and (g) total toxaphene calculated using single and multiple response factors (SRF, MRF). Scales on the right pertain to the open bars. ………………………… 88

Figure II-2. Chromatograms of total toxaphene and Cl-7 to Cl-9 homologue groups in NW Alabama air. Top = standard, middle = soil and bottom = air. Peak 1 = B8-1413 (T2, P26), 2 = B8-1412, 3 = B8-1945, 5 = B8-806/809, 6 = B8-229, 7 = B9-1679 (T12, P50) and 8 = B9-2206. …………………………. 90

Figure II-3. Plots of log P (partial pressure, Pa) versus 1/T (ambient temperature, K) TC = trans-chlordane, CC = cis-chlordane, TN = trans-nonachlor, tox = toxaphene, dieldrin and γ-HCH. The solid line is the linear regression using all data points; the dashed line is the linear regression after 1-2 points (shaded) are removed. ……………………………………………………………. 95 Paper III Figure III-1. HCH concentrations with depth: a) Lake Superior, b) Lake Ontario …………. 116

Figure III-2. Decline of HCHs in Lake Ontario water ……………………………………….. 117 .

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Figure III-3. Enantiomer fractions of chiral OCPs in water and air ……………………. 121

Figure III-4. Plot of α-HCH enantiomer fraction (EF) versus α-HCH air concentrations (pg m-3), dashed lines are the corresponding average EF of α-HCH in water ….. 125

Paper IV Figure IV-1. Cruise track on Lake Superior showing station numbers (Table 1) …………….. 133

Figure IV-2. Chromatograms of total toxaphene and Cl-7 to Cl-9 homolog groups, top = air, middle = water and bottom = standard. Peak 1 = P26, 2 = B8-1412 [22] (no Parlar number), 3 = P39, 4 = P40+P41, 5 = P42, 6 = P44, 7 = P50 and 8 = P63. P26 appears lower than actually present in air and water samples because it splits between silicic acid fractions 1 and 2. Fraction 2 is shown here ……………………………………………… 141

Figure IV-3. Averaged relative proportions of Parlar congeners Peak 3, Peak 5 and Peak 6 normalized to Peak 4 (=1.00) for air and water samples ……………. 141

Figure IV-4. Monthly fugacity ratios (bars), air and water temperatures (solid and dashed lines) for Lake Superior ………………………………………………….. 144

Figure IV-5. Monthly toxaphene fluxes (bars) and wind speed (solid line) for Lake Superior .. 144 Paper V Figure V-1. Sampling and Cruise Track, from BERPAC-93 (—) and Arctic Ocean

Section-94 (---). Small numbers indicate locations of some sampling stations. Extent of ice cover for August 1994 (---) ………………………………………… 152

Figure V-2. Figure 2: Latitudinal trends of HCHs on AOS-94 and BERPAC-93. Bering Sea to the North Pole = increasing numbers; Pole to the Greenland Sea = decreasing numbers.

a) α- and γ-HCH concentrations in water b) α- and γ-HCH concentrations in air c) α/γ-HCH ratio in air and water d) Fugacity ratios of α- and γ-HCH ……………………………………… 155

Figure V-3. Potential and actual net fluxes of α- and γ-HCH at different latitudes. Actual flux = potential flux x fraction of open water. Positive flux = volatilization, negative flux = deposition …………………………………… 164

Figure V-4. Enantiomeric ratios (ERs) of α-HCH in air and water at different latitudes. ER = (+)α-HCH/(–)α-HCH ......................................................................................... 165

Figure V-5. Chromatograms (BSCD column) showing enantioselective degradation of α-HCH with depth at stations AOS-37 and 38 …………………………………. 166

Paper VI Figure VI-1. Map of TNW-99 cruise track …………………………………………………… 174

Figure VI-2. Arctic ice maps, June and August, 1999 ………………………………………… 182

Figure VI-3. Clausius-Clapeyron plots for α- and γ-HCH at RB and TNW-99 ………………. 184

Figure VI-4. EF of α-HCH and concentration of γ-HCH at Resolute ……. 186 Bay (pg m-3). Arrow indicates ice break up.

Figure VI-5. EF in the water versus EF in the air, showing a correlation when >90% open water (r2= 0.68), but no correlation when 0-50% open water …………….. 187

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Paper VII Figure VII-1. Cruise track of AOS-94. Dots running from the Chukchi Sea to the

Greenland Sea correspond to station numbers on Table 1 ………………………… 193

Figure VII-2. Concentration of OCs in the upper 40-60 m of the water column , summarized by latitude (N) HCHs: 65-69 = station 1 + BERPAC-93 data; 70-74 = station 2 + BERPAC-93 data; 75-79 = stations 7,11,13,16; 80-84 = stations 18,19,20,24,25,26; 85-89 = stations 28,29,30,31; 90= station 35; 84-80= stations 37,38; 75 = station 39. Other OCs: 65-69 = station 1; 70-74 = station 2; 75-79 = stations 11,13,16; 80-84 = stations 20,24,25; 85-89 = stations 28,29,31; 90 = station 35; 84-80 = station 37; 75 = station 39. Bar shades are: α-HCH (black) and γ-HCH (white), CHBs: single response factor (black) and multiple response factor (white), heptachlor epoxide (black), trans-chlordane (black) and cis-chlordane (white), endosulfan-I (black) and endosulfan-II (white), trans-nonachlor (black) and cis-nonachlor (white) …………………………………………………………………………… 200

Figure VII-3. Chromatograms of the 7-, 8- and 9-chlorinated CHBs in surface water at station 37 ………………………………………………………………………… 203

Figure VII-4. Enantiomeric ratios (ERs) of α-HCH in the dissolved (–) and particulate (---) Phases ……………………………………………………………………………. 205

Figure VII-5. Chromatograms of heptachlor exo-epoxide (HEPX), cis-chlordane and trans-chlordane enantiomers in the dissolved phase at station 35 ……………… 205

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Chapter 1.0

1.1 Conclusions

Goals of this research were to assess the south-eastern United States as a source of OCPs to the

atmosphere and to determine the state of air-water gas exchange of OCPs in the Great Lakes and Arctic.

Concentrations of chlordanes and toxaphene in air decreased by an order of magnitude between Alabama

and the Great Lakes and underwent a similar decrease between the Great Lakes and the Arctic. Dieldrin levels in

Alabama and Great Lakes air were similar but an order of magnitude lower in the Arctic, while HCH

concentrations in all three regions were fairly uniform. Within the Great Lakes Basin, higher concentrations of

chlordanes and dieldrin were found over the lower lakes of Erie and Ontario, which are nearer to urban and

agricultural sources, than over the upper lakes of Huron and Superior.

Examination of chemical markers offers explanations for these trends. Isomer profiles of toxaphene in

Great Lakes and Alabama air show the similar weathered pattern, indicating that the southern U.S.A. is the source

to the Great Lakes. Toxaphene isomer patterns are different in the Arctic, being more similar to the unweathered

technical standard. This implies a ‘fresh’ source or perhaps a remnant of toxaphene deposited during times of

usage. At the height of usage, toxaphene would have been transported and deposited into the Arctic Ocean and

regional seas where degradation is likely very slow due to cold temperatures. The current atmospheric signature in

arctic air may be a combination of volatilization of unweathered toxaphene from the water and long range transport

from past source regions. Proportions of chlordane enantiomers in Great Lakes air suggest a mix of sources from

termicide usage with racemic patterns and agricultural soil nonracemic patterns, the former dominating in the

southern U.S. and the latter dominating in the Great Lakes Basin. HCHs are quite volatile and have global sources

which resulted in more uniform atmospheric distributions, although occurrence of nonracemic α-HCH in air over

large water bodies including Lake Superior and the Arctic Ocean indicate re-emission contributions to the

boundary layer.

OCP levels in water vary due to differences in transport pathways and the physical characteristics of the

water bodies. The water column of the Arctic Ocean is permanently stratified, with the upper Polar Mixed Layer

(~50 m) consisting of cold low-salinity water from ice melt. This is underlain by water masses of Pacific (middle)

and Atlantic (lower) origins. Within the Arctic Ocean, surface HCH concentrations follow the order: Beaufort Sea

> western Archipelago > Central Arctic Ocean > Bering-Chukchi Seas ~ eastern Archipelago. Water in the

Beaufort Sea was advected from the western regional seas when global emissions of HCHs were highest in the

1970s-early 1980s. An Arctic Ocean Mass Balance Box Model estimated that HCH inputs to the western Arctic

came from a combination of atmospheric deposition and transport by ocean currents (Li et al., 2004). This

advected water is now trapped in the ice-capped Beaufort Gyre and is slowly being released by eastward drainage

through the Canadian Archipelago. Meanwhile, global HCH emissions dropped during the 1980s (Li and

Macdonald, 2005) and during the period of this thesis research HCH concentrations in the Bering and Chukchi

seas were lower than in the Beaufort Gyre. Lower levels of HCHs were also found in the eastern Archipelago

where low-HCH water from the North Atlantic water enters Baffin Bay.

A different spatial pattern was seen for toxaphene, where concentrations in surface water were higher in

the eastern Archipelago and Greenland Sea. This may be due to atmospheric transport from the southern U.S.A. to

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eastern Canada (Ma et al., 2005).

The Great Lakes are a diverse ecosystem, with each lake having physical characteristics different from the

others including seasonal stratification. Highest concentrations of HCHs and toxaphene were found in Lake

Superior, the coldest lake with a low sedimentation rate and the longest water retention time. Lower concentrations

of these OCPs occurred in lakes Erie and Ontario. These are warmer, have faster sedimentation rates and shorter

water retention times. These characteristics allow OCPs to be more quickly removed through water outflow,

sedimentation and degradation. Unlike HCHs and toxaphene, concentrations of chlordanes and dieldrin showed

less variation among the Great Lakes. As noted above, atmospheric concentrations of these OCPs were higher over

the lower lakes. More rapid removal from the lower lakes may be compensated by higher air concentrations and

local sources, resulting in rather uniform water concentrations over all lakes.

To assess the gas exchange state of OCPs in the Great Lakes and Arctic, concentrations in air and water

are required. Although atmospheric monitoring programs are in place for the Great Lakes and the Arctic,

concentrations of most OCPs in water were not well known. Additionally, air measurements from monitoring

programs are usually taken close to the shoreline but are overland whereas paired air and water sampling over

water give the best estimate of gas exchange. The shipboard sampling expeditions conducted in this thesis resulted

in a better understanding of OCPs in air and water of these ecosystems.

The Henry’s law constant (HLC) is also required to estimated the air-water saturation state. HLCs as

functions of temperature were determined by a bubble stripping technique for components of technical chlordane,

p,p’-DDE and hexachlorobenzene (HCB). Temperature dependent HLCs had been previously determined in this

laboratory using the same methodology for toxaphene and HCHs (Jantunen et al., 2000; Sahsuvar et al., 2003).

Based on good agreement with HLCs determined by other methods (toxaphene and HCB) and with a

thermodynamically consistent evaluation from other properties (HCHs, Xiao and Wania, 2004), the HLCs used in

this thesis for these OCPs are considered accurate. The bubble stripping method may produce artificially high

HLCs for more hydrophobic compounds such as chlordanes and p,p'-DDE.

The saturation state of OCPs was calculated for HCHs, toxaphene and cyclodienes in the Great Lakes and

HCHs in the Arctic. The water/air fugacity ratios (FR = fW/fA) varied with compound and location. FR >1 were

found for α-HCH in the Arctic Ocean and toxaphene in Lake Superior, which indicated net volatilization. HCHs

in the Great Lakes were generally near equilibrium (FR ~1), with occasional excursions toward net volatilization or

deposition, and γ-HCH was undergoing net deposition in the Arctic Ocean. An assessment of exchange for

chlordanes in the Great Lakes using the HLCs determined in this work concluded that they were near equilibrium

or volatilizing. However, when the thermodynamically consistent values of Shen and Wania (2005) were used,

chlordanes were undergoing net deposition. Such uncertainties exemplify the need for accurate HLCs.

Estimates of the saturation state from HLCs and concentration data are mathematical predictions utilizing

a fugacity model. This thesis presents direct evidence and confirmation of volatilization, showing that ‘fugacity

works’. FRs predicted the potential for α-HCH to volatilize from the Arctic Ocean. In regions of extensive ice

cover, volatilization was inhibited and air concentrations were lower than in regions of open water where evasion

of α-HCH could occur.

The enantiomers of α-HCH were used as gas exchange tracers to provide further evidence of

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volatilization. In situations of extensive ice cover in the Arctic Ocean, the EFs in boundary-layer air were nearly

racemic, indicating long-range transport from emission regions. In regions of open water or where the ice had

receded, the EFs in the overlying air had the same non racemic enantiomeric signature as the water. This tracer

technique was also used to examine volatilization from Lake Superior, where higher air concentrations were

associated with less racemic more degraded α-HCH.

Increased volatilization of α-HCHs in open water regions has climate change implications for the arctic

and Lake Superior, where increases in temperature are leading to less ice cover and warmer surfaces. This loss of

ice cover will increase the release of HCHs and possibly other OCPs and remobilizing them into the atmosphere

for redistribution. In the case of chemicals which undergo net gas-phase deposition to the Arctic Ocean and Great

Lakes, loss of ice cover will provide more surface area for atmospheric loadings.

There have been long standing questions: ‘can land-based air measurements be used to predict loadings to

a lake or ocean?’ and ‘Are the loadings under- or over-estimated from land based stations?’ Results from this

thesis for α-HCH suggest that concentrations of volatilizing chemicals are likely to be higher in the air boundary

layer over water than at a land station. Thus, air measurements from land may lead to over prediction of fluxes

from the lake.

As a follow-up to the studies in this thesis, sampling was done in Lake Superior in 2005 and the Arctic in

2007-2008. Toxaphene did not decline in Lake Superior between 1996-97 and 2005 (Chapter 6.2.2, Table 5).

Stable concentrations over a decade indicate that removal mechanisms such as outflow, sedimentation and

volatilization are slow and toxaphene is resilient to degradation. As also noted by Swackhamer et al. (1999),

toxaphene is going to be in Lake Superior for multiple decades. HCHs declined by factors of ~2 in Lake Superior

(Chapter 6.2.2, Table 5) and ~3 in the Canadian Archipelago since the 1990s. Outflow is probably not the

dominant removal mechanism of HCHs from Superior, because the concentrations of toxaphene would have

dropped proportionally, knowing tributary inputs are not significant. EFs of α-HCH in Superior declined from

0.450 in 1996-97 to 0.431 in 2005. This indicates that microbial degradation is a significant removal pathway, as

seen in the Arctic.

The consequences of the decades of contaminants studies in the Great Lakes and arctic have already been

seen, the banning of several persistent substances studied here by the Stockholm convention in 2004 and 2009.

Toxaphene and chlordane were on the original list while HCHs were added more recently. Bans or restriction of

organochlorine pesticides have led to reduced levels in the environment. These successes can aid international

regulators in predicting the transport and persistence of ‘new’ chemicals released into the environment.

1.2 Recommendations

Development of alternative methods for determining Henry’s Law Constant for more hydrophobic

chemicals. The use of microporous tubing has been presented in the literature (Xie et al., Atmospheric

Environment, 2004) and may be a viable option but needs further investigations. Determination would start with

hexachlorocyclohexanes, chlordanes, p,p’-DDE and hexachlorobenzene. Additionally, after the method has

proven successful, move on to currently used pesticides, brominated flame retardants and perfluorinated

compounds.

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There has been a long standing question of over-land versus over-water air measurements. Can over land

measurements be extrapolated to predict air-water gas exchange direction and flux magnitude. More paired land

based and over water based air measurements are required. In 2011, there is a Lake Superior Intensive, using

Integrated Atmospheric Deposition Network sites and water surveillance cruises parallel air and water samples can

be taken. This will assess the applicability of land based measurements in predicting gas exchange fluxes.

Further investigation of OCPs in the Arctic Ocean, focusing on the Canadian Archipelago. Concentrations

of OCPs have declined between 1999 and 2007/08. Is this an actual decrease in water concentrations or has there

been a change in water circulation? This will be further investigated in the summer of 2010, when another arctic

cruise is planned.

Investigate transport and fate of currently used pesticides and new and emerging compounds. Endosulfan

was discussed briefly in this thesis, but other currently used pesticides and new and emerging compounds are

transported to the arctic via the atmosphere and by ocean currents. These are under studied compounds with

poorly understood transport mechanisms.

2.0 INTRODUCTION

The organochlorine pesticides (OCPs) investigated in this thesis are on the lists of substances which have

been banned or severely restricted under the United Nations Environmental Program (UNEP) Stockholm Convention

and/or the United Nations Economic Commission for Europe Convention on Long-Range Trans-boundary Air

Pollution (UN-ECE-CLRTAP), due to their persistence, toxicity, bioaccumulation and long range transport potential.

They include toxaphene, components of technical chlordane and hexachlorocyclohexanes (HCHs). Over the past 20

years, environmental concentrations of OCPs in the Great Lakes Basin have generally declined (Sun et al., 2006

a,b; Glassmeyer et al., 1997; De Vault et al., 1996; Hickey et al., 2001) and the Arctic (Braune et al., 2005; Hung

et al., 2005) due to restrictions on usage, but concerns still remain because these substances persist in the

environment and accumulate in the food chain. The purpose of this research is to examine the occurrence of OCPs in

the environment, from a past usage-potential source region in the southern U.S.A. to the Laurentian Great Lakes and

the Arctic. The study employs distinctive proportions of OCP stereoisomers (“chemical markers”) to make inferences

about sources and trace air-water gas exchange.

Since the mid-1980s atmospheric deposition has been recognized as a large, and in some cases dominant,

loading process for persistent organic pollutants (POPs) to the oceans (Duce et al., 1991), the Great Lakes and the

Arctic (Barrie et al., 1992; Hoff et al., 1996; Strachan and Eisenreich, 1988, Swackhamer and Armstrong 1986,

Swackhamer et al., 1999). Large bodies of water such as oceans and lakes play an important role in the global

processes that distribute OCPs, acting either as a sink or a source to the environment. Atmospheric loadings take

place by precipitation, dry deposition of particles and air-water gas exchange. For the chemicals discussed in this

thesis, gas transfer is the dominant atmospheric loading process to large lakes and oceans, whereas dry and wet

deposition are less important. Gas exchange is a ‘two-way street’ and may alternate between net deposition and net

5

volatilization in response to changing temperatures and atmospheric levels of OCPs in the short term and changing

in air and water concentrations over a longer scale. Generally, OCPs are now within a factor of five of equilibrium

but are in a constant state of short term seasonal adjustment (Mackay and Bentzen, 1997). Knowing the air-water

equilibrium status in different regions is important to understanding global source-sink contributions.

This thesis is presented as a collection of papers that have been published in peer-reviewed journals. The

first section is an overview which gives background information and summarizes methods and findings of each

study. This is followed by the papers in their final published format.

CHAPTER 3. COMPOUNDS INVESTIGATED

3.1. Toxaphene

Toxaphene (CAS number 8001-35-2) is a complex mixture, which results from the chlorination of camphene (a

bicyclic terpene derived from α-pinene). The initial reaction forms 2-exo-2,10-dichlorobornane via a Wagner-

Meerwein rearrangement (Parlar, 1985) and completes to yield polychlorinated bornanes. The product contains small

amounts of chlorinated camphenes (Kimmel et al., 2000) and toxaphene is also known as polychlorinated camphenes

(PCC). In the former Soviet Union, polychlorinated terpenes were produced by direct chlorination of α-pinene with

azobis-isobutyronitrile as an initiator (Nikiforov et al., 2004). The theoretical number of different chlorinated

compounds in technical toxaphene is ~16000 (Vetter, 1993), but only ~250 have environmental significance (Hainzl

et al., 1994). The group of toxaphene-like compounds found in the environment has been termed chlorobornanes

(CHBs) and the congener distribution is usually transformed from the technical standard (Muir and de Boer, 1993).

Here, Σtoxaphene is used to represent the sum of quantified components. Toxaphene has an average composition of

C10H10Cl8 (MW 414), with an average chlorine content of 67 - 69%. The seven, eight and nine chlorinated

homologues are the most abundant in the technical mixture.

A B C

Figure 1: Structure of toxaphene: A) B8-1413, B) B8-2229 and C) B9-1679.

Toxaphene was first manufactured by Hercules Co. in 1945, and was registered for use in the U.S in 1947.

Toxaphene was used alone and in combination with other pesticides, such as ethyl and methyl parathion, DDT and

lindane (IARC, 1979; WHO, 1984). Toxaphene was the most popular substitute for DDT after its U.S. ban in 1972.

Total global usage is estimated at 1.33 Mtonnes. The top ten countries using toxaphene and similar products

(Ktonnes) were: U.S.A. (490), former Soviet Union (254), Nicaragua (79), Mexico (71), Egypt (54), Brazil (50),

6

Syria (33), France (26), Colombia (23) and former East Germany (22) (Li and Macdonald, 2005). Peak usage in the

U.S. was in 1974, when two thirds of all production was applied to cotton crops in the south eastern states (Voldner

and Li, 1995). It was also used on vegetables, small grains, soybeans and insect control on livestock. Estimated total

usage in Canada varied between 100-1000 tonnes (Voldner and Li, 1993) and 20-50 tonnes (Muir et al., 2005).

In addition, Canada and the U.S. used toxaphene as a piscicide for non-game fish eradification, chosen over

rotenone because it was cheaper and more lethal (Miskimmin and Schindler, 1994). Canada began phasing out the

use of toxaphene between 1970-1980 and it was deregistered in 1982 (Muir et al. 2005). The U.S. deregistered

toxaphene for most uses in 1982, but stores could be used until 1986. All registered uses of toxaphene mixtures in the

U.S.A. were cancelled in 1990 (ATSDR, 1996). Toxaphene is one of the persistent organic pollutants listed for

elimination of production and use under the Stockholm Convention and UN-ECE-CLRTAP. See Muir et al. (2005)

for more detailed discussions of toxaphene production and usage.

3.1.1. Toxaphene congeners

About 40 individual congeners in the technical toxaphene mixture have been identified and synthesized

(Vetter and Oehme, 2000), examples of structures commonly found in environmental samples are given in Table 1

and Figure 1. Due to the complex nomenclature of toxaphene congeners, several short-hand designations are used.

For example, two persistent and bioaccumulating congeners are an octachloro compound 2-endo,3-exo,5-endo,6-

exo,8,8,10,10-octachlorobornane, also known as Parlar 26, T2 and B8-1413 and a nonachloro compound 2-endo,3-

exo,5-endo,6-exo,8,8,9,10,10-nonachlorobornane, designated as Parlar 50, T12, B9-1679 and Toxicant Ac. The

endo,exo nomenclature usage in Table 1 is consistent with International Union of Pure and Applied Chemistry

(IUPAC) rules (Stern et al., 1992) and differs from some reports (Frenzen et al., 1994). Two common toxaphene

congeners which are produced by degradation of higher chlorinated toxaphenes in sediments are hex-sed (B6-923)

and hep-sed (B7-1001), respectively. The Andrews and Vetter (1995) numbering system is preferred because B-

numbers are available for more congeners than the P-numbers, see Table 1 for nomenclature.

Although peaks matching individual congeners are often quantified as if they are single components, they are

probably not single compounds in air or water samples. Shoeib et al. (1999) showed this by using multidimensional

gas chromatography-electron capture detection (ECD) to examine the composition of toxaphene peaks in air samples

collected on the north shore of Lake Ontario. De Boer et al. (1997) also used multi-dimensional GC-MS and took

heart cuts of technical toxaphene and biological samples to investigate the complexity of the major peaks whose

retention times matched those of B8-1413, B7-515, B9-1679, B9-1025 and B10-1110 standards. All cuts consisted

of 4-12 peaks except B10-1110. Peaks in fish samples showed less complexity in the heart cut analysis,

demonstrating bio-transformation and/or selective uptake.

7

Table 1: Toxaphene Nomenclature

Andrews and Vetter Chemical Name (IUPAC) Parlar a Others b,c

1995

B6-923 2-exo,3-endo,6-exo,8,9,10-hexachlorobornane Hex-Sed

B7-1001 2-endo,3-exo,5-endo,6-exo,8,9,10-HCBd Hep-Sed

B7-515 2,2,5-endo,6-exo,8,9,10-HCB P32 ToxB

B8-1413 2-endo-,3-exo,5-endo,6-exo,8,8,10,10-OCBe P26 T2 and Tox 8

B8-1412 2-endo,3-exo,5-endo,6-exo,8,8,9,10-OCB

B8-531 2,2,3-exo,5-endo,6-exo,8,9,10-OCB P39

B8-1414 2-endo,3-exo,5-endo,6-exo,8,9,10,10-OCB P40

B8-1945 2-exo,3-endo,5-exo,8,9,9,10,10-OCB P41

B8-806/809 2,2,5-endo,6-exo,8,9,9,10-OCB P42a,b ToxA

B8-2229 2-exo,5,5,8,9,9,10,10-OCB P44

B9-1679 2-endo,3-exo,5-endo,6-exo,8,8,9,10,10-NCBf P50 ToxAc, T12 and Tox9

B9-1025 2,2,5,5,8,9,9,10,10-NCB P62

B9-2206 2,exo,3-endo,5-exo,6-exo,8,8,9,10,10-NCB P63

B10-1110 2,2,5,5,6-exo,8,9,9,10,10-decachlorobornane P69

a: Frenzen et al., 1994b: Saleh, 1991c: Stern et al., 1992d: HCB: Heptachlorobornanee: OCB: Octachlorobornanef: NCB: Nonachlorobornane

The most stable structure for toxaphene congeners is the staggered endo-exo-endo-exo conformation of chlorine

atoms on the six-membered ring; e.g. the 2-endo, 3-exo, 5-endo, 6-exo conformation that appears in B7-1000, B7-

1001, B7-1002, B8-1413, B8-1414, , B8-1412 and B9-1679. Vetter and Scherer (1999) found that this structure

confers stability on the toxaphene molecule, rendering it less degradable in the environment and less easily

metabolized. B9-1679 and three octachlorobornanes above have been identified in marine mammals (Stern et al.,

1992; 1996; Vetter et al., 1994; 1997) and B7-1001 of the heptachlorobornanes is one of the dominant congeners

in sediments from toxaphene-treated lakes (Donald et al., 1998; Stern et al., 1996). B8-1413 and B9-1679 were

very slowly, if at all, degraded in treated anaerobic sewage sludge (Vetter et al., 2001) and B9-1679 was not

enantioselectively degraded in biota from the Baltic Sea, the Arctic and Antarctica (Buser and Müller, 1994).

Even though B9-1679 is a nine chlorine substituted congener and can be called ‘bulky’ because of the placement of

the chlorine atoms, it is free of steric hindrance and ring strain. Parlar et al. (2001a,b) divides the toxaphene

congeners into three categories: the first are those with one chlorine atom at each of the secondary ring atoms in an

alternating orientation, corresponds to the endo-exo confirmation above. Examples of this are B8-1413; B8-1414

8

and B9-1679, these are the most stable congeners. The second are those with a geminal dichloro group on the ring

and one chlorine atom in the α-position and includes B8-531 and B9-715. This group makes up about 70% of the

total toxaphene technical mixture and are easily degraded by photolysis or reaction with oxygen species. The third

group is intermediate which is somewhat stable except when a second dichloro group exists at position 10; e.g.,

B8-2229 and B9-1025.

A different approach to ranking the relative environmental/biological stability of toxaphene congeners is

based on their heats of formation, ∆Hf (Vetter and Scherer, 1999). As ring strain and steric hindrance increase so

do the ∆Hf. Lower ∆Hf are associated with more stable structures. The ∆Hf for the above listed stable congeners

ranges from –230 to –245 kJ mol-1. B7-515 and B8-806/808 are less stable in the environment and this is

reflected in their higher ∆Hf (–216 and –216 kJ mol-1). B8-1945 is more stable in the environment and this

reflected in a lower ∆Hf (–237 kJ mol-1). B8-2229 is also a common residue in environmental media, so it is

tempting to say that it is also a stable congener even though it lacks the alternating endo-exo-endo-exo

conformation, The ∆Hf for B8-2229 is –206.78 kJ mol-1, about the same as for less stable congeners. The

prevalence of B8-2229 may be because it is also a degradation product of B9-1025 (Ruppe et al., 2004; Vetter,

1998; Vetter and Scherer 1999).

3.2. Cyclodienes

Cyclodiene pesticides include chlordane and associated compounds, aldrin, dieldrin, endrin, heptachlor,

endosulfan and metabolites heptachlor exo-epoxide and oxychlordane. See Figure 2 for several cyclodienes.

3.2.1. Chlordane

Chlordane (1,2,4,5,6,7,8,8a-octachloro-2,3,3a,4,7,7a-hexahydro-4,7-methanoindene, CAS #: 57-74-9) was

first produced in 1945 and is a mixture of over 140 different compounds (Kirk-Othmer, 1995). It is synthesized by a

Diels-Alder condensation reaction between cyclopentadiene and hexachloro-1,3-cyclopentadiene, forming chlordene

(Sittig, 1980). Technical chlordane is then produced by the addition of two Cl-atoms across the chlordene double

bond at high temperature and pressure (Kirk-Othmer, 1995). The resulting chlorine content ranges from 64-67% with a

purity of 60-75%. The most abundant components in the technical mixture are trans-chlordane, cis-chlordane, trans-

nonachlor, β-chlordene and heptachlor (Buchert et al., 1989; Mattina et al., 1999). Heptachlor exo-epoxide and

oxychlordane are persistent and bioaccumulative metabolites of heptachlor and several chlordane isomers, respectively.

9

Cl

Cl

ClCl

Cl

Cl

Cl

Cl

Cl

ClCl

Cl

Cl

Cl

Cl

Cl

ClClCl

Cl

ClCl

Cl

Cl

ClClCl

Cl

ClCl

Cl

Cl

ClClCl

Cl

Cl

Cl

Cl Cl

Cl

ClClCl

Cl

Cl

Cl

Cl

A B C

D E F

Figure 2: Structure of chlordane A) trans-chlordane, B) cis-chlordane, C) trans-nonachlor, D) oxychlordane, E) dieldrin

and F) heptachlor exo-epoxide.

Technical chlordane was produced since 1948 by Velsicol Chemical Company for usage as an agricultural

pesticide on corn and citrus, for home lawns and gardens and as a termiticide in house foundations. Technical heptachlor

was also used as a termiticide, alone or in combination with technical chlordane. Technical heptachlor contains 22%

trans-chlordane (Kutz et al., 1991). At the height of production in 1971, chlordane was the second-most widely used

organochlorine insecticide in the United States, with the annual production about 11 million kg per year. Over 70,000

tons of chlordane have been manufactured since 1946 (Dearth, 1990). In the mid-1970s, 35% of chlordane was used

by pest control operators, mostly for termites, 28% for agricultural crops, 30% for home lawns and gardens and 7%

for turf and ornamentals (ATSDR, 1994). Between 1983-1988, usage was restricted to only termite control and all

uses were cancelled after 1988. Usage of chlordane in Mexico was stopped in 2003 (NACEC, 2003). A total of 9000

tonnes of chlordane was produced and used in China, the peak year was in 1999 (Liu et al., 2009). Chlordane and

heptachlor are prohibited under the Stockholm Convention and UN-ECE-LRTAP.

3.2.2. Dieldrin

Dieldrin (1,2,3,4,10,10-hexachloro-6,7-epoxy-1,4,4a,5,6,7,8,8a-octahydro-endo,exo-1,4:5,8-

dimethanonaphthalene, CAS #: 60-57-1) and aldrin (1,2,3,4,10,10-hexachloro-1,4,4a,5,8,8a-hexahydro-endo, exo-

1,4:5,8-dimethanonaphthalene, CAS #: 209-00-2) were produced originally in 1948. Aldrin was produced by the

Diels-Alder condensations of hexachloro-1,3-cyclopentadiene with norbornadiene, and dieldrin was produced by

epoxidation of aldrin (ATSDR, 2002).

Cl

Cl

ClCl

Cl

Cl

Cl

O

Cl

Cl

ClCl

Cl

Cl

Cl

O

Cl

Cl

ClCl

Cl

Cl

Cl

O

Cl

Cl

Cl

ClCl

Cl

Cl

Cl

O

Cl

Cl

ClCl

Cl

Cl

Cl

O

Cl

10

Dieldrin enters the environment through direct application or from the use of aldrin, which quickly transforms

into dieldrin in the environment, U.S. production of these two cyclodienes was 90% aldrin and 10% dieldrin

(Jorgenson, 2001). While estimates of aldrin and dieldrin production vary, their production maximized in the mid-

1960s at ~9 million kg/year (Jorgensen, 2001; Kannan, 2005; ATSDR, 2002). Both aldrin and dieldrin were used

heavily in the 1950s to 1970s to combat insects on corn, cotton and citrus crops and also used as termiticides. In the

U.S.A., the peak usage of aldrin was in 1966. Dieldrin and aldrin were de-registered for most purposes in 1987, but

dieldrin remained part of a termiticide formulation until 1990 (ATSDR, 2002). Both compounds are prohibited under

the Stockholm Convention and UN-ECE-LRTAP.

3.3. Hexachlorocyclohexane (HCH)

Technical HCH is a mixture of several isomers. The composition is often reported as: α-HCH (60-70%, CAS

#319-85-6), β-HCH (5-12%, CAS #319-85-7), γ-HCH (10-15%, CAS #58-89-9), δ-HCH (6-10%, CAS #319-86-8) plus

minor isomers (Iwata et al., 1993), however Breivik (1999) quotes larger variations in isomer percentages. The only

insecticidally active isomer is γ-HCH and when produced in 99% purity is known as lindane. See Figure 3 for structures.

HCH was formerly known as benzene hexachloride because it is produced by the photo-chlorination of benzene in the

presence of a free-radical initiator; e.g., visual or UV light, x-ray or γ-rays (Kirk-Othmer 1985).

Figure 3: Structure of HCH

HCH is an insecticide used for agriculture purposes, including fruits, vegetables, seed treatment and forest crops.

It is also used against insects on cattle and as a pharmaceutical to treat lice and scabies. The estimated global application

of technical HCH was 10 million tonnes of between 1948-1997 (Li, 1999). China was the largest producer and consumer

of technical HCH, with 4464 kilotonnes used between 1956-1983. Other heavy-use countries were the former Soviet

Union (1960 kilotonnes, 1950-1990) and India (1057 kilotonnes, 1948-2000) (Li and Macdonald, 2005). European use

of technical HCH was 382 kilotonnes between 1970-1996 (Breivik et al., 1999). Li and Macdonald (2005) show higher

European usage beginning in 1948, with 795 kilotonnes used by France, Spain and former East Germany. The U.S.A.

used 343 kilotonnes of technical HCH from 1948-1977. U.S. production of technical HCH stopped in 1976 but

importation continued from France, Germany, Japan, Spain and China. Over the years, lindane has replaced technical

HCH in most countries. By breaking down total HCH usage by isomer, Breivik et al. (1999) estimated that the

proportion of α-HCH applied in Europe dropped from ~71% in 1970 to ~50% in the mid-1980s to ≤15% in the 1990s.

Technical HCH was deregistered in the U.S.A. in 1978, when it was replaced by lindane. Lindane was deregistered in

Canada in 2004 for all uses except pharmaceutical. Under the terms of the North American Regional Action Plan

(NARAP, 2006) for lindane, the U.S.A. has requested voluntary cancellations of lindane registration and Mexico is

Cl Position

aaeeee αααα-HCH

eeeeee ββββ-HCH

aaaeee γγγγ-HCH

aeeeee δδδδ-HCH

a

e

Cl Position

aaeeee αααα-HCH

eeeeee ββββ-HCH

aaaeee γγγγ-HCH

aeeeee δδδδ-HCH

a

e

a

e

11

working toward a phase-out of lindane for agricultural uses. HCHs are prohibited by UN-ECE-LRTAP and are

candidates for inclusion in the Stockholm Convention.

CHAPTER 4. RELEVANT PROPERTIES AND PROCESSES

4.1. Physicochemical properties

Relevant physical-chemical properties that describe environmental partitioning are saturation vapour pressure

(Psat, Pa), solubility in water (CW,sat, mol m-3) and solubility in octanol (CO,sat, mol m-3 ) (all subcooled liquid

properties) and the three partitioning properties between octanol/water (KOW = CO,sat/SW,sat), octanol/air (KOA =

CO,sat/CA,sat) and air/water (KAW = CA,sat/CW,sat). The concentration of a chemical in air at equilibrium with its liquid is

CA,sat (mol m-3) = Psat/RT. The three partition coefficients are interrelated, KOA = KOW/KAW (Figure 5). Table 2

summarizes properties of the OCPs as the “final adjusted values” (FAVs) of Shen and Wania (2005) and Xiao et al.

(2004) where possible (see Section 4.1.1). Table 2 also includes HLCs measured in this thesis and selected properties

from other reports.

4.1.1 Henry’s Law Constants

The Henry’s Law constant (HLC, Pa m3 mol-1) or air-water partition coefficient describes the ratio of the

partial pressure in air (P, Pa) to the concentration in the water (CW, mol m-3) at equilibrium:

WC

PH = (1)

Figure 4: Triangular Relationship between KOA, KOW and KAW (adapted from Xiao et al., 2004).

octanol dissolved

phase

KOA KAW

KOW

aqueous dissolved

phase

gas phase

12

Table 2: Physical Chemical Properties

HCB TC CC TN p,p'-DDE αααα-HCH γγγγ-HCH Dieldrin Endo-I HEPX Toxaphene Reference

Vapour Pressure 0.094 0.010 0.0073 0.0037 0.0034 0.25 0.076 0.014 0.0044 0.022 1,2

(Pa, at 25oC) 1.15 x 10-5 3

Solubility (mg L -1) 0.40 0.61 0.53 0.25 85 72 5.0 2.6 5.1 1,20.55 3

Log KOW 5.64 6.27 6.2 6.35 6.93 3.94 3.83 5.48 4.94 5.42 1,24.77-6.64 4

Log KOA 7.21 8.83 8.83 9.7 7.46 7.74 8.84 8.49 8.59 1,29.3 5

HLC 0.55 0.24 635 29 27 32 33 0.45 7

(Pa m3 mol-1) 65 6.8 5.7 4.2 0.74 0.31 1.1 0.70 1.7 1,20.67 3

15.9 5.5 10.7 5.1 0.3 0.27 0.99 0.82 2.3 8

TC: trans-chlordane; CC: cis-chlordane; TN: trans-chlordane; HEPX: heptachlor exo-epoxide

1: Shen and Wania, 2005; 2: Xiao et al., 2004; 3: Murphy et al., 1987 (for the technical mixture, a waxy solid with a low and variable melting range); 4: Fisk et al., 1999;5: Harner et al., 2001; 6: Jantunen and Bidleman, 2000; Chapter I; 7: Sahuvar et al., 2003; 8: Cetin et al., 2006

The HLC is affected by concentration, for the relation to be accurate the solution must be dilute. For compounds that

are slightly soluble, the HLC can be predicted by the ratio of the saturation vapour pressure (Psat) and the water

solubility (SW) of the pure organic liquid.

satW,

sat

C

PH = (2)

KAW = H/RT is often called the dimensionless HLC. KAW equals the equilibrium ratio of CA/CW, or when estimated

from saturation properties, KAW = CA,sat/CW,sat, as defined in Section 4.1.

The temperature dependence of H is given

R

S

RT

HHln HH ∆

+∆

−= (3)

where ∆HH and ∆SH are the enthalpy and entropy of the phase change from the dissolved phase to the gas phase and

are assumed to be independent of temperature (Schwarzenbach et al., 1993).

Henry’s Law constants for many OCPs are available in the literature as a function of temperature (Cetin et al.,

2006; Sahsuvar et al., 2003; Staudinger et al., 2001; Paper I), however, literature values often differ substantially.

This could be due to different techniques used in the laboratories which reported the HLCs and/or measurement

artifacts, as discussed in the following section. Although, one could select HLCs from literature compilations (e.g.,

Mackay et al., 2006), a more rigorous approach is to derive probable values of the HLCs (and other properties) by

taking advantage of the relationships among various properties (Figure 4). This was done by Shen and Wania (2005a)

and Xiao et al. (2004) who compiled and evaluated measured properties data from the literature, selected literature-

derived values through averaging or linear regression and made estimates of the uncertainty of these values. These

uncertainty estimates were applied in making relative adjustments to HLC values which were derived from combined

properties, their so-called “final adjusted values” (FAVs).

13

When calculating the HLC for saline waters, an adjustment is made to account for the salting-out effect. This

adjustment increases the HLC value because organic compounds are generally less soluble in seawater (Cetin et al.,

2006; Gossett, 1987; Kucklick et al., 1991; Rice et al., 1997; Schwarzenbach et al., 1993; Staudinger and Roberts,

2001; Xie et al., 1997). For organochlorine compounds studied in this thesis, the factor increase in the HLC ranged

from 1.0 or no effect for γ-HCH to 5.8 for endosulfan-I (Cetin et al., 2006). Cetin et al., (2006) found a relationship

between the molar concentration and the salting out or the Setschenow constant (kS):

Log (HS/H) = kSCS or (4)

Log(S/SS) = kSCS (5)

where HS and SS are the HLC and solubility of the compound in saline water, H and S are the HLC and solubility in

deionized water and CS is the molar concentration of the salt solution. The kS for OCPs ranged from 0.04 for γ-HCH

to 1.8 L mol-1 for endosulfan-II.

4.1.2 Determination of HLCs by the Gas Stripping Method.

The HLC was determined by the inert gas stripping (“bubble stripping”) technique, pioneered by Mackay et al.,

(1979). Kucklick et al. (1991) describes this method for HCHs, see Paper I for details. The inert gas stripping apparatus

is a three-chamber nested vessel (Figure 5). The outer chamber contains air for insulation, the next chamber contains

water circulated from a water bath at a controlled temperature and the inner chamber, where the gas stripping takes place,

contains the aqueous solution of the compounds. The volume and height of the inner chamber are 525 mL and 62 cm.

The purge tube, with a coarse frit on the end, is lowered to a few millimeters from the bottom of the inner chamber.

Kucklick et al., (1991) showed that the air bubbles were in equilibrium with HCHs in water by sampling at two different

depths (45-50 and 26 cm). The top of the inner chamber narrows and coils three times then leads to the outlet, this coil is

to prevent aerosols from escaping the system.

HLC experiments are run in the stripping or dynamic head space mode. In the former, the decrease in water

concentration produced by gas stripping is followed over time and the HLC is calculated according to equations in Paper

1. The dynamic head space method employs shorter stripping times and the effluent compound vapours are collected in

an adsorbent trap at the outlet of the apparatus. Analysis of the trap contents and the water yields air and water

concentrations, from which the HLC is calculated (Sahsuvar et al., 2003).

For less water soluble compounds there may be a bias in determining the HLC by the inert gas stripping

method which leads to anomalously high values. Lei et al. (2006) observed this bias in the HLCs for normal alkanols.

They compared the inert gas stripping results to those determined using a static head space sampling method and

HLCs calculated from literature values of Psat and CW,sat. The inert gas stripping technique over-estimated the HLC

for longer chain alkanols at lower temperatures, while shorter chain alkanols and measurements done at higher

temperatures showed no artifact. The artifact is caused by adsorption of the more hydrophobic compounds at the

bubble-water interface. They quantified this surface artifact using the enhancement factor, where an enhancement

factor of 1 indicates no bubble adsorption artifact. The enhancement factor increased with chain length of the

alkanols, at lower temperatures and higher flow rates, all resulting from adsorption to the bubble surface.

14

N2

air insulation

constanttemperature

airoutlet

water outlet

bubblingchamber

Constant Temperature water bath

pre-saturator column

Figure 5: Henry’s Law constant apparatus.

Shunthirasingham et al. (2007) estimated a positive enhancement factor for some of the OCPs. For trans-

chlordane, cis-chlordane and p,p’-DDE they predicted surface artifacts between 5-40oC. They predicted a surface

artifact for HCHs only at lower temperatures, while no artifact was predicted for HCB. Comparing the HLCs

determined experimentally in this thesis and other studies with the FAVs (Shen and Wania, 2005; Xiao et al., 2004)

α-HCH shows no bias for all published values, where γ-HCH shows a bias below ~20oC for Cetin et al. (2006) and

Kucklick et al. (1991) but not for other authors. At 5oC, the enhancement factor for γ-HCH ranges for ~2.0-3.5. The

magnitude of the enhancement factor depends on the size of the molecule, for example HCB and α-HCH did not show

a bias but the larger molecules trans-chlordane (enhancement factor ~3-6 at 25oC), cis-chlordane (~2-6) and p,p’-

DDE (~2-8) did, and the bias increased with decreasing temperature.

Comparing the HLCs presented in Paper I with the FAVs reported by Shen and Wania (2005), the result for

HCB determined in this study is about a factor of two lower than the FAV value (35 vs. 65 Pa m3 mol-1) while the

results for cyclodienes and p,p’-DDE were 5-8 times higher. It is likely that bubble adsorption artifacts account for the

difference between my HLC and the FAV values for trans-chlordane and cis-chlordane but not DDE. As a gauge of

hydrophobicity, the KOW value, relative to HCB, for TC is 5.0 times higher, 4.0 times for CC and 20 times for DDE.

The FAV for α-HCH are similar those determine by inert gas stripping studies (Cetin et al., 2006; Jantunen and

Bidleman, 2001; Kucklick et al., 1991; Sashuvar et al., 2003;) but differ for γ-HCH in Cetin et al. (2006) and

Kucklick et al. (1991).

4.2. Air-water gas exchange

15

4.2.1. Fugacity and the net exchange direction

Oceans, seas and lakes play an important role in the global cycling of OCPs, either acting as a sink or a source

for OCPs in the environment. The direction of diffusive exchange of OCPs across the air-water interface may be

altered by air and water temperature and concentrations, and also by ice cover that limits the air-water exchange.

Exchange of gases between water and air occurs continuously in an attempt to attain equilibrium. The approach to

equilibrium can be assessed by comparing fugacities of the chemical in water and air (fW, fA) (Wania et al., 1998).

The fugacities are related to the concentrations in air (CA)and water (CW) by the following equations:

fW = CWH (6)

fA = CARTA (7)

AA

W

A

W

RTC

HC

f

fFR == (8)

where CW and CA are the dissolved and gaseous concentrations in water and air (mol m-3), H is the Henry's law

constant at the temperature of the water (Pa m3 mol-1 ), R is the gas constant (8.314 Pa m3 mol-1 K-1) and TA is the

temperature of the air (K). Fugacity ratios (FR = fW/fA, eq 9) of <1.0 and >1.0 imply net deposition and volatilization

respectively, and FR = 1.0 indicates air-water equilibrium.

4.2.2 Uncertainty in Fugacity Ratios

Uncertainty in the air-sea exchange calculations arise from systematic and random errors, which can be substantial.

Fugacity ratios (FR) will vary in the short term to due changes in atmospheric concentrations resulting from shifting

wind patterns (e.g., from clean areas to over cities) and diurnal cycles (Lee et al., 1998). Several authors have

discussed ways of estimating the random errors in FRs (Bruhn et al., 2003; Hillery et al., 1998; Hoff, 1994; Hoff et

al., 1996; Mackay and Bentzen, 1997). Due to relatively large uncertainties in the terms involved in calculating FRs,

the deviation from equilibrium must be large to unequivocally indicate a net flux direction. Errors in the Henry’s law

constant are often the limiting factor and tightening the precision of HLCs is a good way to improve precision in the

FRs. To minimize uncertainties in the concentration terms, it is preferable that parallel air and water measurements be

made and the analysis done in the same laboratory.

Uncertainties in FR values in Papers III-VI were estimated by propagation of errors in CA, CW and the Henry's law

constants (Paper I; Sahsuvar et al., 2003). The error propagation equation used was (Paper III):

2H

2Ca

2Cw

2FR RSDRSDRSDRSD ++= (9)

4.2.3. Rate of gas exchange: the modified two film model

Air-water gas exchange either over lakes or oceans has been discussed by many authors (Ballschmiter 1992;

Bidleman and McConnell, 1995; Breivik et al., 2002; Eisenreich et al., 1997; Hornbuckle et al., 1994; Mackay 1991;

16

Meng et al., 2007; 2008; Odabasi et al., 2008; Pacyna et al., 1998, Wurl et al., 2006; Papers III-VI) but all use the

same two-film approach when estimating net gas exchange direction.

This two film gas exchange model, first developed by Whitman (1923) and modified by Liss and Slater (1974)

and Mackay and Yuen (1983), can be used to estimate the net gas flux (N, mol m-2 d-1). This model assumes the

mass transfer is limited by diffusive exchange across the air and water films at the interface. Written in fugacity

terminology:

Deposition: NDEP (mol m-2d-1) = DAW (fA) (10)

Volatilization: NVOL (mol m-2d-1) = DAW (-fW ) (11)

Net flux: NNET (mol m-2d-1) = DAW (fA- fW) (12)

These equations are set up such that deposition is positive and volatilization is negative, a sign convention that is

consistent with usage by the IADN program (Blanchard et al., 2008) and used in Papers III and VI of this thesis. The

opposite convention (volatilization positive, deposition negative) was common in some older literature (e.g., McConnell

et al., 1996; Ridal et al., 1996) and was used in Papers IV and V of this thesis. When the fugacity ratios and fluxes

indicate equilibrium or no net flux, there is still volatilization and deposition occurring but at the same rate. For this

reason, it is important to present the separate fluxes due to volatilization and deposition rather than only the net flux

(Murphy et al., 1987). In eq (10-11):

RT

K )Padm (molD OG1-1-2-

AW = (13)

and WAOG RTk

H

k

1

K

1 += (14)

and

61.0

21/33/1A

0.5

101-

A)19.7 + V(

1/29) + (1/M 0.3) + U15(0.2 =)s (cm k

Σ (15)

and 3.0

M1.6410

1-W 29.6

V U0.45 =)h (cm k

−

(16)

In eq 13, DAW is the overall mass transfer coefficient expressed in fugacity terms, KOG is the overall mass transfer

coefficient expressed on a gas-side basis (m d-1) and kA and kW are individual mass transfer coefficients for the air and

water films. The mass transfer coefficients were calculated using eq 15 from Galarneau et al. (2000) which were

simplified from Hornbuckle et al. (1994). In equations 4 and 5, ΣVA is the sum of atomic diffusion volumes, calculated

from the incremental volumes in Table 11-1 of Mackay and Yuen (1983), VM is the molar volume ( cm3 mol-1) and U10 is

the wind speed at 10 m (m s-1 ).

Mass transfer coefficients can also be calculated from relationships in Mackay and Yuen (1983).

where kA (m s-1 ) = 10-3 +46.2 x 10-5 (6.1 + 0.63 U10)0.5 U10 ScA

-0.67 (17)

and kW (m s-1) = 10-6 +3.41 x 10-5 (6.1 + 0.63 U10)0.5 U10 ScW

-0.5 (18)

17

where ScW and ScA are the Schmidt numbers for the compounds in water and air (Bidleman and McConnell, 1995) which

physically relates the relative thickness of the hydrodynamic layer and mass-transfer boundary layer. Schmidt

numbers for the OCPs were estimated from those of PCB congeners having similar molecular weights (Bidleman and

McConnell, 1995). When these two sets of equations were used to estimate mass transfer coefficients for toxaphene, the

values were 40-60% larger for kW and 2-5% smaller for kA. Since the Henry’s Law constants for toxaphene and HCHs

are relatively low compared to other OCPs, over 95% of the resistance to transfer lies in the air film and KOG is dominated

by kA (eq 14). Thus, the KOG values for toxaphene calculated by the two approaches differ by <10%. The discrepancy

is slightly higher for chlordanes, averaging ~15% and slightly lower for HCHs, averaging ~5%.

4.3. Chemical tracers of exchange processes

4.3.1. Isomers and parent-metabolite pairs

Ratios of isomers (e.g., trans-chlordane/cis-chlordane) or parent/metabolite ratios (e.g., DDT/DDE) are often

used to make inferences about chemical age and sources. However, the compounds have different physicochemical

properties and reaction rates in the atmosphere, which lead to changes in proportions during volatilization and long-

range atmospheric transport. For example, the ratio of p,p’-DDT to p,p’-DDE (DDT/DDE) has been used to infer the

age of residues, where a larger ratio for DDT/DDE implies a fresher source and a smaller ratio implies an older

source. However, the liquid-phase vapour pressure of p,p’-DDE is 6.8 times higher than for than p,p’-DDT at 20oC

(Hinckley et al., 1990) so the relative ratio of the parent to metabolite changes when the two compounds undergo

volatilization from soil (Kurt-Karakus et al., 2006; Liu et al., 2009).

Another example of this is for HCHs, technical HCH contains α-, β- and γ-HCH isomers. HLCs are in the

order α-HCH > γ-HCH > β-HCH, so the three isomers are removed from the atmosphere at different rates by

precipitation and air-to-water gas deposition. Thus, long range transport over the oceans results in discrimination

among isomers (Iwata et al., 1993b; Li et al., 2002). γ-HCH also undergoes photochemical degradation at a faster rate

than the α-isomer (Brubaker and Hites, 1998). Observed higher ratios of α-HCH/γ-HCH in the air of remote regions

could be due to technical HCH input or more efficient transport of α-HCH.

Technical chlordane also contains many components, including octa- and nona-chlorinated compounds.

Ratios of trans-chlordane, cis-chlordane and trans-nonachlor are used to infer sources; e.g., emissions from relatively

unweathered chlordane from house foundations where it was applied for termite control versus chlordane in

agricultural soils where there is a greater chance for microbial degradation and dissipation (Eitzer et al., 2001).

Because these chlordane components have different vapour pressures (Hinckley et al., 1990), fractionation takes place

when they volatilize from soils and this confounds the comparison of source and atmospheric chlordane proportions.

Also, seasonal differences in the trans-chlordane/cis-chlordane ratio have been found in air, higher in the summer

months than the winter. It is hypothesized that trans-chlordane is less stable photochemically (Bidleman et al., 2002;

Gouin et al., 2007; Hoff et al., 1992; Hung et al., 2005), but evidence is lacking.

4.3.2. Chiral Compounds

Chiral compounds have two or more enantiomers, which exist as non-superimposable mirror images. When

manufactured, chiral chemicals often contain identical amounts of the two enantiomers, also called a racemic mixture.

18

Enantiomers have identical physical-chemical properties and abiotic degradation rates, but behave differently in a

chiral environment. Enantiomers can have different biotic degradation rates because enzymes are also chiral

molecules (Buser and Müller, 1992). They also elicit different toxicological responses. For example, the (–)

enantiomer of o,p'-DDT is a more active estrogen mimic than the (+) enantiomer (Hoekstra et al., 2001; McBlain,

1976). Enantioselective degradation of a chiral compound can occur in soil and water as a result of microbial activity,

which leads to depletion of one enantiomer. This altered enantiomer signature can provide information on the

transport and fate of OCPs (Section 6.2). The enantiomer ratio (ER) and fraction (EF) are defined by the quantities of

the (+) and (–) enantiomers.

)(

)( ER

−+= (19)

)( )(

)(

1ER

ER EF

−+++=

+= (20)

The EF is preferred over the ER because ER boundaries are 0 to infinity where EF is bounded by 0 and 1 (Harner et

al., 2000; Ulrich et al. 2003). ER values lead to skewed data distributions and statistical summaries, such as mean and

standard deviation of ER values can be misleading. Chiral compounds discussed in this thesis are: α-HCH, trans-

chlordane, cis-chlordane and the metabolites heptachlor exo-epoxide and oxychlordane.

4.3.3. α-HCH Enantiomers

α-HCH exists as two enantiomers, see Figure 6. Enantioselective processing of α-HCH takes place in water,

soils and biota.

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl Cl

ClClCl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl Cl

ClCl

Figure 6: Enantiomers of α-HCH.

Enantioselective degradation of (+)α-HCH seems especially pronounced in cold oligotrophic water systems

such as arctic and subarctic lakes and wetlands (Falconer, et al., 1995a; Helm et al., 2000; Law et al., 2000), the

Arctic Ocean (Paper VII; Falconer et al., 1995b, Harner et al., 1999; Moisey 2001) and the upper Great Lakes

(Huron, Superior) (Paper III; Law et al., 2000). The lower Great Lakes (Erie, Ontario) also contain nonracemic α-

HCH residues, but the extent of enantioselective degradation is less than in lakes Superior and Huron (Paper III, Ridal

et al., 1997). Small temperate lakes contain α-HCH residues that are generally racemic (Law et al., 2001). A trend

has been noted for increased enantioselective degradation of α-HCH in lakes with longer water residence times (Law

et al., 2001). Preferential degradation of (+)α-HCH has also been found in North Carolina estuaries (Venkatraman et

al., 2003), the North Sea (Faller et al., 1991; Hühnerfuss et al., 1992), the Kattegat Sea off the Swedish west coast

(Sundqvist et al., 2004) and the Baltic Sea (Wiberg et al., 2001).

19

Not all aquatic systems show depletion of (+)α-HCH; preferential degradation of (–)α-HCH was observed in

surface water of the Bering and Chukchi seas (Papers V and VI), some regions of the North Sea (Faller et al., 1991)

and in passive air samples from British Columbia on the west coast of Canada (Shen et al., 2004). Hoekstra et al.

(2003d) found near racemic α-HCH in water on the north shore of Alaska near Barrow. Preferential depletion of (–

)α-HCH in North Pacific air below the marine boundary layer was contrasted to near racemic α-HCH in air that was

in the free troposphere and less likely to have equilibrated with the ocean (Genualdi et al., 2009). Law et al., (2004)

found that microbial degradation of α-HCH, with preferential loss of the (–) enantiomer, occurred in groundwater at a

contaminated pesticide reformulating and packaging facility in Florida.

Enantiomeric analysis has been used to assess bioaccumulation and biotransformation of α-HCH in

marine food webs (Fisk et al., 2002; Hoekstra et al., 2003d; Moisey et al., 2001; Wiberg et al., 1998; 2000). Marine

mammals have the capability to enantioselectively biotransform α-HCH but the observed enantioselectivity varies by

organ and species and is not predictable. (+)α-HCH is preferentially accumulated in the brain of seals and eider

ducks, while (–) α-HCH is often absent (Kallenborn and Hühnerfuss, 2001) suggesting active uptake. (+)α-HCH can

more easily penetrate the blood-brain barrier than (–)α-HCH (Möller et al., 1993; Ulrich et al., 2001), the brain

barrier acts as a chiral guard, stereochemically separating the enantiomers. The higher concentrations of α-HCH in

the local environment may induce different levels of enzymatic activity, leading to greater shifts from racemic

(Kallenborn and Hühnerfuss, 2002). Additionally, the health state of the animal may change the metabolic capacity to

degrade of α-HCH (Kallenborn and Hühnerfuss, 2002; Wiberg et al., 1998). Fisk et al. (2002) determined α-HCH

enantiomers in ringed seals from the Northwater Polynya in eastern Canada and found racemic values, whereas slight

depletion of (+)α-HCH was found in water, zooplankton and arctic cod, suggesting there was some degradation of the

(–) enantiomer by seals. Hoekstra et al. (2003d) found racemic residues in seawater and arctic cod, but depletion of

the (–) enantiomer in bowhead whale, beluga and ring seal, and depletion of the (+) enantiomer in bearded seal.

Wiberg et al. (2000) found racemic α-HCH in arctic cod. Blubber from harbour and grey seals showed depletion of

the (–) enantiomer (Klobes et al., 1998) and small cetaceans showed depletion of the (–) enantiomer, although this

varied among species (Tanabe et al., 1996). Blubber may not reflect seals’ capacity to metabolize α-HCH, as Wiberg

et al. (1998; 2000) found nearly racemic α-HCH in blubber but non-racemic α-HCH in liver. Fisk et al. (2003) found

a negative trend of EFs in ringed seals with age, although Wiberg et al., (2000) found no trend with age or sex. Wong

et al. (2002) and Warner et al. (2006) demonstrated that rainbow trout and zooplankton did not enantioselectively

degrade α-HCH in laboratory feeding experiments.

Identifying the factors that influence enantioselective processes in soils and water is crucial in understanding

the enantiomer composition of residues. Although racemic α-HCH is commonly found in soils (Aigner et al., 1998;

Falconer et al., 1997; Kurt-Karakus et al. 2005; Li et al., 2006; Wiberg et al., 2001), enantioselective degradation

also occurs (Falconer et al., 1997; Kobličková et al. 2008; Kurt-Karakus et al. 2005, 2007; Li et al., 2006; Meijer et

al. 2003; Shen et al., 2009). Investigation of factors suggest that enantioselective degradation is related to the humic

and fulvic acids, organic carbon, total nitrogen and higher clay content, these all encourage the growth of soil

microflora (Kobličková et al. 2008). Suar et al. (2005) found that gene strains linA1 and linA2 in the soil bacterium

(Sphingomonas paucimobilis) degraded α-HCH enantioselectively. A1 specifically metabolized the (+) while A2

20

metabolized the (–)-enantiomer but degradation of α-HCH by the entire bacterium was not enantioselective. Buser

and Müller (1995) found the removal rate of (+)α-HCH exceeded that of the (–) enantiomer in bench-scale studies

using anoxic sewage sludge.

4.3.4. Cyclodiene Enantiomers

Many components of technical chlordane are chiral (Karlsson et al., 1999), including those examined in this

thesis work: trans-chlordane, cis-chlordane, and metabolites heptachlor exo-epoxide and oxychlordane (Figure 2).

Enantiomers of trans-chlordane are shown in Figure 7.

Cl

Cl

ClClCl

Cl

Cl

Cl

Cl

Cl

ClCl Cl

Cl

Cl

Cl

Cl

Cl

ClClCl

Cl

Cl

Cl

Cl

Cl

ClCl Cl

Cl

Cl

Cl

Figure 7: Enantiomers of trans-chlordane

The enantiomer distributions of chiral cyclodienes have been determined in many types of media from

temperate North America and the Arctic: Agricultural soil (Aigner et al., 1998; Eitzer et al., 2001; 2003; Falconer

1997; Leone et al., 2001; Wiberg et al., 1991), anaerobic marine sediments (Li et al., 2007), ambient air in temperate

and tropical regions (Bidleman et al., 1998a,b; Daly et al., 2007; Eitzer et al., 2003; Gouin et al., 2007; Leone et al.,

2001; Shen and Wania, 2005b; Ulrich and Hites, 1998; Venier and Hites, 2007; Wong et al., 2008, Papers II and III),

indoor air (Leone 2000, Paper II), arctic air (Bidleman et al., 2002, 2004), Great Lakes water (Paper III); Arctic

Ocean water (Paper VII) and arctic biota (Borgå and Bidleman, 2005; Fisk et al., 2001, 2002a,b; Hoekstra et al.,

2003b,d; Wiberg et al., 2000). Non-racemic patterns are generally seen in these media, with the exception of

chlordanes in indoor air and soil around houses treated with technical chlordane as a termiticide (Eitzer et al., 2001;

Leone et al., 2000; Paper II) and Arctic Ocean water (Paper VII; Hoekstra et al., 2003d) and some low tropic level

arctic biota (zooplankton, char and cod) (Borgå and Bidleman, 2005; Hoekstra et al., 2003d).

Like α-HCH, the EFs of chlordanes provide information on the bio-transformation capability of a species and

the trophic transfer of contaminants in the food chain (Wiberg et al. 2000; AMAP 2004). Several studies have shown

the ability of organisms to biotransform or eliminate chiral chlordanes enantioselectively (Kallenborn and Hühnerfuss,

2001; Warner and Wong 2006; Wong et al., 2002) and enantioselective metabolism or transport of chlordanes has

also been demonstrated in plants (Lee et al., 2003; Mattina et al., 2002, 2004, 2006).

Non-racemic chlordane residues in arctic biota have been reported in ringed seals (Fisk et al. 2002; Hoekstra

2003d; Wiberg et al. 2000); bearded seals (Hoekstra et al., 2003d); polar bears (Wiberg et al., 2000); bowhead and

beluga whales (Hoekstra et al., 2003d) and sea birds (Fisk et al., 2001; Ross et al., 2008). Seabirds had varying

patterns of depletion of enantiomers suggesting different mechanisms for chlordane metabolism among species.

Additionally, EFs did not predict concentrations nor trophic level (Fisk et al., 2001). Fisk et al. (2002) found that EFs

of chlordanes in ringed seals varied with δ13C and δ14N which indicate trophic level, suggesting that diet and exposure

may play a role in EFs observed. Ringed seal food generally had racemic chlordane residues but residues in ringed

21

seals were non-racemic, implying that seals have the capacity to enantioselectively degrade chlordane (Hoekstra et al.,

2003d). Metabolites of chlordane, oxychlordane and heptachlor exo-epoxide were also non-racemic in ringed seals

(Fisk et al. 2002; 2000; Hoekstra et al., 2003d; Klobes et al. 1998; Wiberg et al. 1998), bearded seals, bowhead and

beluga whales (Hoekstra et al., 2003d); polar bear (Wiberg et al. 2000), seabirds (Fisk et al., 2001; Ross et al., 2008)

and arctic cod (Moisey et al., 2001; Wiberg et al., 2000). Warner and Wong (2006) and Wong et al. (2002)

investigated the uptake and elimination of chlordanes by aquatic invertebrates and rainbow trout and found that both

uptake and metabolize chlordane enantioselectively, additionally the metabolite, oxychlordane, was produced.

4.3.5. Enantiomers in source identification and exchange processes

When chiral chemicals volatilize from soil and water surfaces, they retain their distinctive enantiomer

proportions (Bidleman and Falconer, 1999). Thus, appearance of enantio-enriched chemicals in air indicates

microbially degraded residues that have been recycled to the atmosphere from soil or water. An enantiomer

composition close to the racemic proportion in the original technical product indicates fresh inputs or residues that

have not been subjected to microbial degradation. Chiral α-HCH undergoes enantioselective degradation in

oceans, lakes and wetlands usually with depletion of the (+), but occasionally of the (–) enantiomer (Section 4.3.3).

Volatilization of nonracemic α-HCH from water bodies can be distinguished from transport of racemic α-HCH in

background air.

Soils are of special interest because of their ability to both supply and release POPs to the atmosphere on

a global scale, in a process known as “grasshoppering” (Gouin et al., 2004; Wania and Mackay, 1996). Racemic

trans-chlordane and cis-chlordane were found in soil near house foundations that were treated with technical

chlordane for termite control (Eitzer et al., 2001). Leone et al. (2000) also found high concentrations of racemic

chlordanes in the indoor air of homes in the midwest U.S.A. Agricultural soils generally contain nonracemic

residues, depleted in (+) trans-chlordane and (–) cis-chlordane, and enriched in (+) heptachlor exo-epoxide (Aigner

et al., 1998; Bidleman et al., 2003a; Eitzer et al., 2001; Meijer et al., 2003; Wiberg et al., 2001), although racemic

chlordanes have been occasionally reported (Falconer, 1997; Finizio et al., 1998). Air samples taken directly

above agricultural soil that has a reservoir of these chiral cyclodienes show the same enantiomer pattern as the soil;

this implies that the residues in soil are a source to the atmosphere (Bidleman et al., 1998b; Eitzer et al., 2003;

Finizio et al., 1998; Leone et al., 2001; Meijer et al., 2003). Background soils contain chlordane residues that are

generally non-racemic. Kurt-Karakus et al. (2005) determined the enantiomer composition of chlordanes in

background soils worldwide and found the following frequencies of enantioselective degradation:

Table 3: Distribution of OCPs EFs in background soils, % of total samplesa.

(+) Depletion (–) Depletion Racemic N

trans-chlordane 72 10 17 58

cis-chlordane 28 57 15 54

a) From Kurt-Karakus et al., 2005

22

As with agricultural soils, the dominant depletions in background soils were (+)trans-chlordane and (–) cis-

chlordane, although reversed degradation patterns occurred in some soils, especially for cis-chlordane.

Enantiomer proportions of trans-chlordane and cis-chlordane in arctic air have changed over time from

nearly racemic in the early 1970s to nonracemic in the mid-late 1990s. This implies that the sources of chlordane to

the atmosphere may have changed over time or that the residues in agricultural and background soils have

undergone more extensive enantioselective degradation than in the earlier years (Bidleman et al., 2002, 2004a). A

similar trend has been noted for trans-chlordane in a lake sediment core, collected on Devon Island, NU, Canadian

Arctic (Stern et al., 2005). Trans-chlordane was close to racemic in sediment layers from the late 1940s and 1950s

(EFs 0.490-0.495) and decreased into the late 1990s.

Enantiomers were employed as chemical markers in this thesis to make inferences about sources and air-

water exchange of α-HCH and the chiral chlordanes. Further information and specific applications are discussed

in Section 6.

CHAPTER 5. MATERIALS AND METHODS

5.1 Sampling Locations

Measurements of OCPs in air were made in Muscle Shoals Alabama at the Tennessee Valley Authority

reservation on a year long campaign during 1996-1997 (Paper II). Air samples were collected during five cruises on

the Great Lakes during 1996-2000, Superior, Huron, Erie and Ontario (Figure 8). Lakes Superior and Huron were

covered in August 1996 and May 1997, Lake Erie in August, 1996 and Lake Ontario and the upper St. Lawrence

River in July and September 1998 and June 2000. In addition, four air samples were taken from a buoy in the west

end of Lake Ontario from June to September, 1998. Details can be found in Papers III and IV. Arctic – subarctic

measurements were made during the summer of 1993 as part of BERPAC-93 on board the R/V Okean and in 1994 as

part of AOS-94 on board the CCGS Louis S. St. Laurent, see Figure 8 for cruise tracks. Air samples were taken

continuously in 1999 during Tundra Northwest (TNW-99) on board the CCGS Louis S. St. Laurent and on land at

Resolute Bay (RB) from June to August 1999 (Figure 8). Details can be found in Papers V-VII.

5.2 Air Sampling

Air samples were taken at a number of locations and years, but all employed similar methods. The

particulate and gaseous phase OCPs were collected on a glass fibre filter (Whatman, Maidstone, England, 20 x 25 cm,

EPM 2000, collects 99% of particles >0.3µm) followed by plugs of polyurethane foam (PUF, Olympic Products

Corp, Greensboro, NC, U.S.A.), respectively. Dimensions of the collection media differed among the campaigns; see

descriptions and methods for cleaning before use in Papers II-VII.

23

Figure 8: Sample collection sites, A) Arctic sites: BERPAC-93, AOS-94, TNW-99 and B) Great Lakes sites: 1996-

2000.

Lake Superior August 1996

May 1997

Lake Ontario July/Sept 1998

June 2000

Lake HuronAugust 1996, May 1997

Lake Superior August 1996

May 1997

Lake Ontario July/Sept 1998

June 2000

Lake HuronAugust 1996, May 1997

North Pole

AOS’94BERPAC’93TNW’99

Resolute BayNorth Pole

AOS’94BERPAC’93TNW’99

Resolute Bay

24

5.3 Water Sampling

Low volume (LV, 4L) and high volume (HV, 80-200L) water samples were collected in stainless steel cans,

via a submersible pump or polytetrafluoroethylene lined Go-Flo sampling bottles (General Oceanics). The LV

samples were analysed for HCHs only; the HV samples were analysed for the other OCPs. The water was passed

through a glass fibre filter (baked at 400oC, GF/F, nominal cutoff 0.7 µm Whatman, Maidstone, England) to collect

the particulate OCPs, or centrifuged to remove the particulate matter (Great Lakes work only), and the dissolved

OCPs were concentrated on a solid adsorbent. The LV method employed commercially available ENV+

(polystyrene-divinyl benzene, 30mL cartridge containing 200 mg, Jones Chromatography) or C8-bonded silica

adsorbent cartridges (30mL cartridge containing 1-g, Mega Bond-Elut, Varian) and HV method used glass columns

hand-packed with XAD-2 (polystyrene-divinyl benzene) resin (75 mL 20-60 mesh, Supelco). Multiple LV and HV

samples were taken in lakes Ontario and Superior, and LV samples were taken in Lakes Huron and Erie. Depth

profiles were done in lakes Superior and Ontario. During BERPAC-93, AOS-94 and TNW-99, LV and HV surface

water samples were taken. Cruise tracks are shown in Figure 8. Samples at depth were taken on BERPAC-93 and

AOS-94; Descriptions of the collection media, methods for cleaning before use, and station locations are given in

Papers III-IV for the Great Lakes and V-VII for the Arctic.

5.4 Sample Extraction and Cleanup/Fractionation

Much of these details were removed from the papers due journal restrictions on length. C8 cartridges for LV

water sampling were extracted with 12 mL DCM (Omnisolv, EMD, Gibbstown, NJ, U.S.A.) ENV+ cartridges were

extracted with DCM/acetone (Omnisolv, EMD, Gibbstown, NJ, U.S.A.). The XAD-2 was placed in a glass column

(100 mL, 3.0 cm i.d.) and 350 mL DCM was slowed dripped through and dried over sodium sulfate. Extracts were

reduced in volume and solvent exchanged into isooctane by rotary evaporation and a gentle stream of nitrogen.

LV samples were cleaned up using a 1-g alumina column (baked at 400oC, 6% deactivated with HPLC

grade water, 70-230 mesh size, EMD, Gibbstown, NJ, U.S.A.) topped with sodium sulfate. The column was eluted

with 12mL 10% DCM/petroleum ether (PE, Omnisolv, EMD, Gibbstown, NJ, U.S.A.). Extracts of HV water and

air samples were cleaned up and fractionated on a column containing 3 grams silicic acid (baked at 400oC,

deactivated with 3% water 100-200 microns, Mallinckrodt Baker INC. Phillipsburg NJ U.S.A.), topped with 1 g

alumina and 1 cm of sodium sulfate (baked at 400oC, granular, EMD). The first fraction was eluted with 30 mL PE

and contained polychlorinated biphenyls and naphthalenes (not analysed here), HCB, heptachlor, p,p’-DDE and

some toxaphene congeners. Fraction two was eluted with 30 mL DCM and contained the rest of the OCPs.

5.5 Analysis Method, Quantitative and Chiral

Samples were analysed by gas chromatography - mass spectrometry operating in electron capture negative

ion mode. For quantitative analysis, DB-5 or DB-5MS capillary columns were used. For chiral analysis several

columns were used, BetaDEX-120, BGB-172 and BSCD (similar to BGB-172 and used in earlier work) and

Restek β-DEXcst. Column dimensions and temperature programming rates varied according the study and are

given in the original papers, as is source information.

25

5.6 Quality Control

Blanks were done for both air and water during each sampling campaign and followed the same procedure

throughout the years and sampling campaigns. Air blanks were done by loading a PUF and filter in the sampling

head and drawing air for 30s. Water blanks were done by passing ~100 mL through the solid phase sorbent.

Travel and lab blanks were also done. Most blanks were below detection limits, so a method detection limit was

calculated based on the lowest standard injected and sample volume.

Spike methods have evolved over the years because 13C and deuterium labelled compounds became more

available in later studies. In earlier studies, samples recovery checks for water were done by spiking lake or ocean

water with a known amount of the compounds of interest, or by adding a low-abundance chemical such as δ-HCH

to each sample. More recently labelled compounds were acquired: α-HCH-d6, γ-HCH-d6, p,p’-DDT-d8, p,p’-DDE-

d8, 13C10 trans-nonachlor, 13C12 dieldrin and 13C10 heptachlor exo-epoxide. Labelled surrogates were added to each

sample so an individual recovery factor could be applied to each samples.

Quality control measures for quantitative analysis: a calibration plot was prepared by injecting a range of

standards that spanned the concentration range of the sample extracts, generally sub-pg uL-1 to ~100 pg uL-1. Two

ions (target and qualifier) for each compound were monitored by the mass spectrometer and the ratio of

target/qualifier was required to be within ± 20% of the standard ratio to ensure peak purity. Although

quantification was done using Chemstation software, the integration of each peak was checked and manually re-

integrated if required.

Quality control for enantiomeric analysis was done by repeatedly injecting racemic standards to determine

the reproducibility of measuring enantiomeric fractions. This process was done for each column and repeated

frequently. The criterion used for peak purity in samples was agreement of the target/qualifying ion ratio within ±

5% of the standard values (Falconer et al. 1997). Confirmation of the enantiomeric analysis was done where

possible by analysis on several columns with differing chiral stationary phases and, for some compounds, differing

enantiomer elution orders.

Chapter 6. SOURCES, TRANSPORT AND ENVIRONMENTAL OCCURRENCE

6.1. Southern Sources and Transport

6.1.1. Toxaphene

Toxaphene was heavily used in the southern U.S. and emission of toxaphene from soil residues is a major

source to the rest of North America (Li, 2001, Li et al., 2001). The sources, air concentrations and atmospheric

transport of toxaphene from the southern U.S. and occurrence in Great Lakes air, water, sediments and biota has

been reviewed by Muir et al. (2005). Concentrations of Σtoxaphene in southern U.S.A. air before deregistration of

toxaphene typically reached hundreds of ng m-3 (Muir et al., 2005). After the final deregistration in 1986,

atmospheric levels dropped to tens to hundreds of pg m-3, although low ng m-3 concentrations were still reported in

some locations (Bidleman et al., 1998c; Hoh and Hites, 2004; James and Hites, 2002; Muir et al., 2005).

Concentrations in tree bark declined quickly with distance from the source (Mcdonald and Hites, 2003).

26

Atmospheric levels of Σtoxaphene in the southern U.S.A. are ~10 higher than in the Great Lakes Basin (Papers II

and IV).

Similar amounts of Toxaphene was used in Mexico and Nicaragua (Section 3.1) and was manufactured at a

plant in Nicaragua until ~1993. Nicaragua heavily used toxaphene on cotton, with soil residues ranged from 17-44

µg g-1 (Carvalho et al., 2003). The Σtoxaphene at 15 air sampling stations in Mexico ranged from 6-689 pg m-3 in

2000-2003 (Alegria et al., 2006, 2008; Wong et al., 2008, 2009). Atmospheric Σtoxaphene concentrations in

Belize averaged 36 pg m-3 in 1995-1996 (Alegria et al., 2000).

Atmospheric transport from the southern U.S. to the Great Lakes and farther north has been well

documented (Hoff et al., 1992a, 1993; MacLeod et al., 2002; Voldner and Schroeder, 1989). Sampling a year

before toxaphene deregistration, Rice et al. (1986) measured a strong gradient in air concentrations from the south

to north, with highest concentrations of 6600 pg m-3 in Mississippi, 1300 pg m-3 in Missouri and 94-360 pg m-3 in

Michigan. More recently, James and Hites (2002) and Hoh and Hites (2004) measured toxaphene in air on a

transect from the southern U.S. to Michigan. They found highest levels in Arkansas (averaging 1400 – 1600 pg m-

3 in 2000-2001 and 2002-2003), followed by Texas (280 pg m-3, 2000-2001) and lower levels at a coastal

Louisiana site (61 pg m-3, 2002-2003), Indiana (60 pg m-3, 2003-2004) and Michigan (10-23 pg m-3, 2000-2003).

Hoff et al. (1992a; 1993) noted that higher levels of Σtoxaphene in southern Ontario occurred when air trajectories

tracked back to the southern U.S. and northern Mexico.

Volatilization of toxaphene from soil residues in the southern U.S. is a major and continuing source of

emissions (Bidleman and Leone, 2004a,b; Li et al., 2001; Ma et al., 2005a; MacLeod et al., 2002). Modelling by

Ma et al. (2005a,b) and MacLeod et al. (2002) indicates that soil emissions followed by episodic transport delivers

toxaphene to the Great Lakes, where it is deposited by wet and dry deposition. Based on current and historical air

concentration data, James and Hites (2002) estimated the time for 50% decrease of atmospheric toxaphene to be 3-

4 years in southern U.S.A. and 4-10 years in the Great Lakes region.

The concentration of Σtoxaphene in Alabama air ranged from 6-611 pg m-3 (Paper II). The annual mean

concentration in Alabama air (176±151 pg m-3) was similar to the mean for July 1994 - January 1995 in Columbia,

South Carolina, 189±107 pg m-3 (Bidleman et al., 1998). The Σtoxaphene in Alabama air was 10-20 times higher than

in Great Lakes air, see next section for details (Paper IV; Glassmeyer et al., 1999; Shoeib et al., 1999).

When, air samples were collected over a period of one year in Alabama, Σtoxaphene concentrations showed

seasonal cycling (Paper II). This was also seen in South Carolina (Bidleman et al., 1998) and over Lake Superior

(Paper IV; Glassmeyer et al., 1999; Shoeib et al., 1999). The relationship of atmospheric toxaphene to air

temperature was investigated by plotting Log P/Pa (partial pressure, Pa) versus 1/Ta. The Alabama data showed a

significant regression (p <0.01) with a slope of -3350. Other studies also showed this relationship over Lake

Superior (-3291, Paper III), in South Carolina (-2583, Bidleman et al., 1998), at Eagle Harbor (-2438, Glassmeyer

et al., 1999), at Point Petre (-2284, Shoeib et al., 1999) and over Lakes Superior and Michigan (-3989 James et al.,

2001) and along transect of the mid-U.S.A. (-3453 Michigan, -4447 Indiana, -3924 Arkansas and -1622 Louisanna,

Hoh and Hites, 2004). Although the toxaphene concentrations in these locations differed by up to 50-fold, the

slopes are quite similar.

27

In Paper II, two persistent toxaphene congeners (B8-1413 and B9-1679) were quantified in ambient air

samples from Alabama. These were the only two congeners quantified because standards of other individual

congeners were not yet readily available at that time. Qualitatively, the chromatographic patterns of toxaphene in

the air more closely resembled those in soil (Bidleman and Leone, 2004a; Harner et al., 1999) than in the technical

toxaphene mixture, implying that the toxaphene in ambient air was probably due to volatilization of aged residues

in the soil (Figure 13) not from current usage. The air and soil both showed transformations from the technical

mixture in the Cl-8 and Cl-9 homologs. Chromatograms of the Cl-8 homolog show a set of four peaks that elute

close together on a non-polar column (Figure 9) and have similar vapour pressures (Bidleman and Leone, 2004a)

but show different proportions in the air and soil compared to technical toxaphene (Bidleman and Leone, 2004a;

Harner et al., 1999). This is due to differences in their stabilities (see Section 3.1.1). Major components of these

four peaks are B8-531, B8-1414/1945 (nearly coelute), B8-806/809 (coelute) and B8-2229 (Papers II and IV).

B8-1414 has the stable endo-exo-endo-exo structure, where B8-2229 does not. B8-2229 is a degradation product

of B9-1025, contributing to it pseudo persistence, as mentioned above. B8-531 and B8-806/809, lacking the endo-

exo-endo-exo conformation, are less stable and are depleted in soils and air (Bidleman and Leone, 2004a; Harner

et al., 1999) probably due to preferential degradation.

6.1.2. Cyclodienes

Chlordane was used for ~40 years mostly as an insecticide on crops, home lawns and gardens, and as a

termiticide. About 80 million Americans live in homes that were treated with cyclodiene pesticides, of that ~65%

were treated with technical chlordane to control termites (ATSDR, 1994). Studies of indoor air in the U.S.A. have

shown that homes built during the 1940-50s have higher levels of chlordane in the indoor air (Offenberg et al., 2004).

Indoor air samples collected in the U.S. cornbelt (Leone et al., 2000) and Muscle Shoals, Alabama (Paper II)

contained much higher levels of chlordanes than outdoor samples, and the composition of chiral chlordanes was

racemic in both studies. Additionally, transformations of parent components in the indoor environment are limited,

since high levels of parent components (trans-chlordane, cis-chlordane and trans-nonachlor) versus low occurrence of

metabolites oxychlordane and heptachlor exo-epoxide were found in indoor air and soils from treated house

foundations (Eitzer et al., 2001; Leone et al., 2000).

Potential sources of chlordanes in ambient air include volatilization from local and regional soils, release from

homes treated for termite control and long-range transport. Shen and Wania (2005b) found higher levels of

chlordane-related compounds and dieldrin in the mid-Atlantic and southeastern U.S.A. compared to other sites in

North America. Offenberg et al., (2004a,b) reported elevated chlordanes in cities of New Jersey, Texas and

California. Park et al. (2001) reported a mean total chlordane of 93 pg m-3 in Texas. Hoh and Hites (2004) reported

highest average concentrations of chlordane in air at a rural site in Arkansas (200 pg m-3) and an urban site in Indiana

(210 pg m-3). Lower concentrations were found at a coastal Louisiana site (59 pg m-3) and at a rural Integrated

Atmospheric Deposition Network (IADN) site in Michigan (39 pg m-3). Higher levels of chlordane in the southern

and eastern U.S.A. and in urban air can be attributed to termiticide usage by compound profiles which are more

similar to those of unaged technical chlordane than to soil residues. Regarding sources to the Great Lakes, Hafner and

Hites (2003; 2005) identified a broad southern region for chlordane, whereas dieldrin sources appeared more

28

centralized in the midwest.

Figure 9: Total toxaphene and homologs Cl-7 to Cl-9 chromatograms showing air and soil from Alabama compared to

a standard, see Paper II for peak labelling.

Chlordane levels in air of southern Mexico in 2000 (21-485 pg m-3, Alegria et al., 2006) and Belize in 1995

(23-257 pg m-3, Alegria et al., 2000) were on the same order of those in the southern U.S., but more recent

measurements in Mexico (2002-2006, 1-18 pg m-3, Wong et al., 2008, 2009) and Costa Rica (2004, 0.7-35 pg m-3,

Daly et al., 2007) were lower and near or below those in the Great Lakes region (Paper III and Table 4). Persistence

in tropical regions are lower due to faster degradation, leading to a increasing gradient from the tropics to the arctic.

In this study (Paper II), elevated chlordane levels were found in homes of Muscle Shoals, Alabama, and levels of

chlordanes in ambient air were 10-20 times higher than over the Great Lakes (Paper III). Concentrations of dieldrin and

heptachlor exo-epoxide were similar in the two regions. Regressions of log p vs. 1/T for these cyclodienes were

significant (p <0.01) with a slopes of –2964 to –4007.

The environmental persistence of chlordane in soils has been well established (Aigner et al., 1998; Bidleman et

29

al., 1998a; Falconer et al., 1997; Finizio et al., 1998; Iwata et al., 1995; Mattina et al., 1999; Ulrich et al, 1998).

Estimates of half life in soils range from less than 10 years to greater than 20 years (Mattina et al., 1999; Printup,

1991; Stewart 1975). Transformations, whether biotic or abiotic, change the relative proportions of applied technical

chlordane components over time. Chlordane also continues to volatilize from soils years after the initial application

thus increasing the ambient atmospheric levels (Eitzer et al., 2003; Finizio et al., 1998). Leone et al. (2001)

demonstrated that soils in the midwestern U.S.A. “Corn Belt” are emitting chlordanes to the atmosphere. Heptachlor

exo-epoxide is produced from the metabolism of heptachlor in soils (Bidleman et al., 1998b) and oxychlordane also

has a source in soils (Falconer et al., 1997; Suprock et al., 1980).

In contrast to soils in the midwest U.S.A., the ΣCHLOR in soils in the southern U.S.A. are an order of

magnitude lower (Bidleman and Leone, 2004b) and it appears that most of the chlordane usage in the south was for

termite control. Models of soil-air exchange found that chlordanes in southern U.S.A. air were not the result of

volatilization from agricultural soils but mainly due to regional transport, probably from volatilization from

termiticide treated house foundations (Harner et al., 2001; Scholtz and Bidleman, 2006). These could include

southern cities, but also many small towns and farms where buildings were treated (Harner et al., 2001; Scholtz and

Bidleman, 2006). Moreover, chlordanes in ambient air of Alabama were close to racemic (Paper II), whereas they

were nonracemic in agricultural soils of the southern U.S.A. (Wiberg et al., 2001).

6.1.3. HCHs

α-HCH was not determined in Alabama samples because of substantial breakthrough problems (Paper II).

Other groups have reported α-HCH in Texas (8-174 pg m-3; Park et al., 2001) and Costa Rica (2-12 pg m-3; Daly et

al., 2007). Levels of γ-HCH in Alabama (mean 50 pg m-3, Paper II) were similar to those in the Great Lakes Basin

(Paper III and Table 4). While larger ranges of γ-HCH were found in Texas (38-403 pg m-3) (Park et al., 2001),

Costa Rica (1.7-170 pg m-3) (Daly et al., 2007) and Mexico (8.8-104 pg m-3) (Wong et al., 2008).

6.2. Great Lakes

6.2.1. Air

Evidence from peat cores in the Great Lakes region and western Canada, and sediment from Lake Ontario and

Siskiwit Lake on Isle Royale in Lake Superior indicate that atmospheric deposition of toxaphene and other OCPs

peaked in the 1960s to 1970s and declined into the 1980s (Donald et al., 1998; Howdeshall and Hites, 1996; Rapaport

and Eisenreich, 1986; Swackhamer et al., 1999). Sediments from Lake Superior itself show a varied picture, with

peak accumulation years ranging from 1969-1991 (Muir et al., 2005).

In Paper IV, the Σtoxaphene levels in air over Lake Superior were higher in August than May, averaging 28

± 10 pg m-3 and 12 ± 4.6 pg m-3 (Table 4). Similar concentrations were found by James et al. (2001) above lakes

Superior and Michigan in the spring and summer of 1997-1998 of 3-54 and 3-57 pg m-3, respectively. James et al.

(2001) also reported levels of Σtoxaphene at Sleeping Bear Dunes on the shore of Lake Michigan of 19-70 pg m-3.

Lower annual toxaphene concentrations were found at Eagle Harbor on Lake Superior in 1996-97 (average 6.6 pg m-3)

(Glassmeyer et al., 1999) and Point Petre on Lake Ontario in 1995-1997 (average 3.8 pg m-3) (Shoeib et al., 1999).

30

Chromatograms of toxaphene in Great Lakes air showed transformations from the technical toxaphene

standard where the air was dominated by the lighter, earlier eluting congeners. Similar profiles were found in

Alabama air (Paper II). Recalcitrant toxaphene congeners B8-1413 and B9-1679 were enriched in the Lake Superior

air samples compared to the technical mixture in both seasons (Paper IV) and Shoeib et al., (1999) found similar

results at Point Petre, Lake Ontario.

Toxaphene in air showed a season cycling (Log P/Pa vs 1/T) with higher concentration in the summer time,

this was found in this study (Paper IV), over Lakes Superior and Michigan (James et al., 2001), at Eagle Harbor

(Glassmeyer et al., 1999) and at Point Petre (Shoeib et al., 1999).

Paper III discusses atmospheric chlordanes in the Great Lakes. Generally, the Σchlordanes (trans- + cis-

chlordane + trans-nonachlor) was lower over Lake Superior (5-13 pg m-3), followed by Lake Huron (6-21) and

higher concentration over lakes Ontario and Erie (12-108 pg m-3) (Table 4). During one sampling campaign on Lake

Ontario, higher concentrations of Σchlordanes were found at the west end of the lake near the Greater Toronto area

(Paper III), these elevated Σchlordanes have also been reported in Toronto (Table 4) and Chicago, and are probably

related to former termiticide usage (Gouin et al., 2007; Harner et al., 2004; Motelay-Massei et al., 2005; Shen and

Wania, 2005b). Gouin et al. (2007) deployed passive samples around the Great Lakes and found decreasing

gradients of chlordane compounds from urban to rural/remote sites.

During the Lake Michigan Mass Balance study in 1995, air samples for trans-nonachlor were collected.

Agricultural regions had higher air concentrations of trans-nonachlor than rural sites. The agricultural samples were

taken in the cornbelt where technical chlordane was applied in the past, and persistent residues in the soil were probably

contributing to the levels in the air. The urban centre of Chicago had the second highest concentration of trans-

nonachlor and over-lake air samples in the southern end of the lake, closer to Chicago were higher compared to the

northern part of the lake (Miller et al., 2001).

Air concentrations of α-HCH for an individual lake tended to be higher in the summer months (July,

August) than in spring (May, June) and fall (September) (Paper III). Considering these measurements and others

(Table 4), average concentrations of α-HCH across the basin spanned a fairly narrow range of 22-81 pg m-3, whereas

the range for γ-HCH was larger (9-165 pg m-3). During the years of this study, lindane was still being used in

Canada, with the heaviest application occurring in the Prairies as a seed treatment for canola and to a lesser extent in

Ontario and Quebec on seed corn (Li et al., 2004; Ma et al., 2003). A coupled atmospheric transport – air/surface

exchange model showed that lindane was transported from the Prairies and deposited into the Great Lakes Basin (Ma

et al., 2003, 2004).

Long term studies at IADN stations identified a decreasing temporal trend in air for α-HCH, γ-HCH, trans-

chlordane, cis-chlordane and trans-nonachlor, heptachlor exo-epoxide and dieldrin. Between 1996-2003, α-HCH

had the shortest half life ranging from 3.1-4.2 years, where others had half-lives ranging from 6.1-13 y, except for

Brule River on Lake Superior where there was no significant trend for trans-chlordane, trans-nonachlor and

heptachlor exo-epoxide (Sun et al., 2006a). The trend in precipitation was not so clear, α-HCH and most of the

cyclodiene pesticides at the IADN sites showed no change between 1996-2003, although γ-HCH did show declining

trend at most sites (Sun et al., 2006b).

31

On the basis of compound profiles, chlordane residues over the Great Lakes were attributed to regional transport

of chlordane emissions from soils; e.g., from former use on corn in the states adjacent to the Great Lakes and regional or

long range transport from past termiticide usage areas in Great Lakes cities and the southern U.S.A. Periodic spikes in

chlordane and dieldrin concentrations in air over the Great Lakes may be due to transport from these source (Sun et al.,

2006a).

The ratio of trans-chlordane/cis-chlordane and trans-chlordane/trans-nonachlor in air over the Great Lakes

were lower in the summer-early fall sampling campaigns than in the spring (Paper III). This result of higher ratios in

the winter and lower in the summer have been found in many other studies in the Great Lakes and Arctic (Bidleman

et al., 2002; Gouin et al. 2007; Halsall et al., 1998; Hoff et al., 1992b; Hung et al., 2002; Oehme et al., 1996; Patton

et al., 1991; Su et al., 2008). Reasons for this are not known, but cyclodienes undergo photodegradation in the

atmosphere (Hühnerfuss et al., 2005) so it has been speculated that trans-chlordane is more labile to photochemical

reaction and therefore more readily lost in the season with more sunlight (Halsall et al., 1998; Oehme, 1991; Patton

et al., 1991). Su et al. (2008) found higher levels of degradation products oxychlordane and heptachlor exo-epoxide

during the summertime, although the latter may be from degradation in soil rather than air (Bidleman et al., 1998b).

Eitzer et al. (2001) found faster degradation of trans-chlordane than cis-chlordane in soils, but Yamada et al. (2008)

found similar photo-degradation rates for trans-chlordane and cis-chlordane. Another factor may be seasonally

differing contributions of chlordane sources to the atmosphere; e.g. termiticide and soil emissions, and volatilization

from the lakes and oceans. Higher ratios of trans-chlordane/cis-chlordane and trans-chlordane/trans-nonachlor due

to termiticide usage have been reported for indoor air (Leone et al., 2000) and urban air in Toronto and Chicago

(Gouin et al., 2007; Shen and Wania, 2005b).

6.2.2. Water

Lake Superior has the highest concentration of Σtoxaphene in water of all the Great Lakes, followed by

intermediate levels in lakes Michigan and Huron and lowest levels in lakes Erie and Ontario (Table 5). Swackhamer

et al. (1999) suggest this is due to the colder water temperatures, longer water residence time and slow sedimentation

rate compared to the other Great Lakes. Swackhamer and Symonik (2004) found seasonal differences in water

concentrations in lakes Superior and Michigan, where concentrations in the epilimnon (above the thermocline) were

lower under stratified conditions but higher when the lakes were unstratified. This may have resulted from

toxaphene volatilizing from the lake when stratified. Although no historical trends of toxaphene in water of the

Great Lakes are available, Glassmeyer et al. (1997) reported that Σtoxaphene did not decline significantly in Lake

Superior lake trout between 1977-1992, although levels declined in fish from the other Great Lakes with half lives of

1.4-5 years. Other studies do show a drop in fish concentrations between 1970-1998 in all the Great Lakes (De Vault

et al., 1996; Hickey et al., 2001). In 1999, 69% of Canadian fish consumption advisories for Lake Superior were

due to toxaphene (Ontario Ministry of the Environment, 1999), although this dropped in more recent guidelines due

to polychlorinated dibenzo-p-dioxins and dibenzofurans being added to the list of contaminants.

In Paper III, concentrations of chlordanes, dieldrin and heptachlor exo-epoxide were determined in surface

water of lakes Superior and Ontario (Table 5). Concentrations of the cyclodienes were only slightly higher in Lake

Superior compared to Lake Ontario. Historical trends show that the concentrations of chlordanes in water have not

32

changed in the connecting channels of Niagara-on-the-Lake and Fort Erie between 1986-1998 whereas heptachlor

exo-epoxide and dieldrin have dropped by ~70% (Marvin et al., 2004; Williams et al., 2000, 2003). Levels of

dieldrin and chlordanes in Great Lakes fish have dropped between 1988/1991 and 1998 in all lakes except Lake Erie

(De Vault et al., 1996; Hickey et al., 2001). In this study, no spatial differences in any OCPs were found among the

stations in individual lakes (Paper III). In the 1990s, higher levels of chlordanes were seen at Niagara-on-the-Lake

than Fort Erie (Niagara River Secretariat, 2000) additionally dissolved concentrations of trans-nonachlor in water

were higher in the south-western part of Lake Michigan near Chicago than the northern basin and deep water samples

had greater concentrations than the shallow samples (Miller et al., 2001).

Like toxaphene, HCHs showed a trend by lake, with the highest concentration in Lake Superior followed by

Huron and lower but similar concentrations for Erie, Ontario and Michigan (Table 5). Historical trends of HCHs in

water are available for most of the Great Lakes and in some of the connecting channels between 1986-1997 or 2001,

they all show dramatic reductions in both α- and γ-HCH (Paper III; L’Italien et al., 2000; Marvin et al., 2004; Williams

et al., 2001; 2003; 2004). Lake Ontario has the longest historical record, where levels of α-HCH in the 1980s average

~4300 pg L-1, and steadily declined, averaging 350 pg L-1 in 2000 (Paper III).

6.2.3. Air-Water Exchange

In Paper IV, the gas exchange of toxaphene in Lake Superior was investigated by making parallel air and

water measurements in August 1996 and May 1997. Chromatographic profiles of toxaphene residues in Lake

Superior water and air were compared to the technical toxaphene standard and it was noted that the air and water

contained a higher proportion of the lighter congeners. Compared to B8-1414 + B8-1945 (co-eluted in this study),

both the water and air showed depletion of the B8-531 and B8-806/809. Toxaphene in the air over Lake Superior

appeared to be due to a combination of volatilization and long range transport; this is supported by the water/air

fugacity ratios that show toxaphene was volatilizing from Lake Superior.

Swackhamer et al. (1999) modelled the historical trend of Σtoxaphene in the Great Lake atmosphere, using

the 1989 atmospheric measurements in southern Ontario of Hoff et al. (1993) and assuming that earlier air

concentrations were proportional to toxaphene production. Based on this trend, the estimated net air-water gas

exchange direction to Lake Superior was depositional throughout the 1970s and 1980s, but the net flux direction was

reversed to volatilization in the 1990s as atmospheric concentrations decreased. A budget for the mid-1990s

indicated that net gas exchange accounted for 70% of annual Σtoxaphene losses from Lake Superior (Swackhamer

1999). James et al. (2001) calculated Σtoxaphene fluxes from air and water measurements during 1997-1998 and

estimated that toxaphene was volatilizing from Lakes Michigan and Superior during 1997-1998 and also predicted

that both lakes would be out-gassing toxaphene for considerable years to come.

In Papers I and III, the gas exchange of chlordanes in the Great Lakes was investigated by making parallel air

and water measurements between 1996-2000. Depending upon which Henry’s Law constants (see section 4.1.1)

were used to calculate air-water exchange, the trans- and cis-chlordanes were at equilibrium or undergoing net

volatilization from the lakes. Heptachlor exo-epoxide and dieldrin showed variations from deposition to

volatilization with season and lake.

FRs for the HCHs indicated near-equilibrium conditions, generally within a factor of 2, with slight excursions

33

toward net volatilization or deposition. Volatilization of α-HCH was indicated from Lake Superior, May 1997 and

Lake Huron, August 1996, and deposition in Lake Ontario, July 1998. Deposition of γ-HCH was found in Lake

Superior in both seasons and in Lake Ontario, July 1998 and June 2000. Volatilization of γ-HCH took place from Lake

Huron, August 1996 (Paper III).

Ridal et al. (1996) collected air and water samples in Lake Ontario in 1993 to study the seasonal changes in the air-water

gas exchange direction for HCHs. They found that, on average, α- and γ-HCH underwent net volatilization in the

summer months but net deposition during the spring and fall. Paper III describes a later sampling campaign on Lake

Ontario where α- and γ-HCH were found to be near equilibrium or undergoing net deposition, even in the summer.

Differences between this study in 1998-2000 and the 1993 one of Ridal et al. (1996) were due to changes in air and

water concentrations in the intervening years. Perlinger et al. (2005) used a diffusion denuder to sample α-HCH in Lake

Superior in summer 2002 and spring 2003 and found that the net flux of α-HCH was out of the lake during the summer

but depositional in the spring. They also found a concentration gradient, with higher concentrations at 1 m compared to

8.5 m above the water.

IADN recently published the loadings report for the years 2000-2005 (Blanchard et al., 2008). Gas exchange

loading estimates of legacy OCPs continued to decline with very small net air-water fluxes for most compounds, <2

ng m-2d-1. The exceptions were dieldrin and HCHs, where dieldrin was volatilizing from all the Great Lakes with a net

flux ranging from –14 to–65 ng m-2 day-1 and was higher than other OCPs except α-HCH in lakes Superior and Huron.

α-HCH was undergoing net volatilization from lakes Superior and Michigan and was close to equilibrium in the

other lakes. Deposition of γ-HCH has decreased between 2000-2005 in the Great Lakes Basin probably in response to

the chemical being banned in 2002-2004 in Canada and 2006 in the U.S.A. (Blanchard et al., 2008). IADN also

addressed the loadings of OCPs from urban centres. Chicago contributed between 2-20% of the total loadings to all of

Lake Michigan for the chlordanes (8.2-19%), dieldrin (7.9-16%) and ΣHCHs (2-4%). The urban contribution was

higher for PCBs and PAHs, 25-50% and 100-330% respectively (Blanchard et al., 2008).

6.2.4. Chiral Tracers of Gas Exchange

Chiral compounds allow us to derive additional information on compound origin and air-water gas

exchange processes. Apportionment of the airborne chiral chemical between two sources categories can be done

knowing the enantiomer fraction of the chemical in source A (EFA), source B (EFB) and in the air sample, resulting

from mixing of these sources (EFM) (Harner et al., 2000):

)EF - (EF

)EF - (EF f

BA

BMA = (21)

A similar equation, based on ER instead of EF, was derived by Bidleman and Falconer (1999). This generally works

best when the chemical in background air is close to racemic; i.e., EFB ~0.5. This situation might occur if source B is

from current use of a racemic pesticide (e.g., long-range air transport of racemic α-HCH in Asian countries) or from

air transport of an aged pesticide residue, but one which has not been subjected to microbial degradation (e.g.,

racemic chlordanes from former termiticide use). Meijer et al. (2003) used this source apportionment model to verify

the equilibrium condition of chlordanes in air sampled directly over agricultural soil.

Depletion of (+)α-HCH was found in surface water of the Great Lakes, with mean EFs: Superior (0.45) <

34

Huron (0.46) = Erie (0.46) < Ontario (0.47) (Paper III). Ridal et al. (1997) found that when fugacity ratios (Section

4.2.1) predicted net volatilization of α-HCH out of Lake Ontario, the enantiomer profile (given as ER in that paper) in

overlying air was similar to that in the water, but when α-HCH was depositing, ERs in air over the lake were closer to

racemic or reflected background air. Bidleman and Falconer (1999) used the Ridal et al. (2007) data and a source

apportionment equation based on ER to estimate that 58% of the α-HCH over Lake Ontario in the summer of 1993

was volatilized from the lake. Shen et al. (2004) found higher atmospheric levels of α-HCH on the shores of Lake

Superior in 2000-2001 compared to continentally North American sites, probably as a result of volatilization from the

lake.

Studies reported in Paper III indicate that, when there was a stable air boundary layer over Lake Superior in

late summer, the EFs in the water and air were closely matched. In this situation, 90-94% of the α-HCH in air over the

lake had volatilized from the lake, as estimated from eq 23. When the boundary layer was weaker in spring, the α-

HCH in air was closer to racemic than in water. Figure 10 shows the relationship between the EF and concentration

of α-HCH in air over Lake Superior. Higher concentrations were associated with more non-racemic residues. This

indicated that α-HCH had volatilized from the lake and increased the concentration in overlying air. Shen et al.

(2004) also found levels of α-HCH were higher and more non-racemic on the shore of Lake Superior compared to

inland continental sites. Similarly, Bethan et al. (2001) found that (+)α-HCH was depleted in air and precipitation on

the North Sea coast during late summer – early fall, but residues were closer to racemic in the colder months.

Figure 10: α-HCH in air: EF versus air concentration for Lake Superior, August 1996 and May 1997.

Lake Superior

0.4450.4500.4550.4600.4650.4700.4750.4800.485

0 50 100 150 200

αααα -HCH Concentration (pg/m 3)

αα αα-H

CH

EF

Lake Superior

0.4450.4500.4550.4600.4650.4700.4750.4800.485

0 50 100 150 200

αααα -HCH Concentration (pg/m 3)

αα αα-H

CH

EF

35

Table 4: Atmospheric Measurements in the Great Lakes and the Arctic (pg m-3)

Location Year αααα-HCH γγγγ-HCH TC CC TN HEPX DIEL Toxaphene ReferenceGreat Lakes

SuperiorAug 1996 and May 1997 75 19 5.7 4.3 3.3 8.6 24 10-41 1,2May 2000-May 2001 81 23 2.1 1.9 0.7 3.1 1.9 3,41996-2003 annual 79 18 2.7 3.3 2.3 5.5 9.0 5July 2002 - June 2003 73 20 3.3 4.6 4.6 - 18 62002-2003 summer 6-170 71996-1998 annual ΣHCH 93 ΣChlordane 8.2 8.6 8Eagle Habor, Sep 96-Dec 97 0-63 9Summers of 1997 and 1998 15 10

HuronAug 1996 and May 1997 54 21 9.5 6.2 5.6 6.6 21 1May 2000-May 2001 37 21 1.2 1.0 3.6 2.1 6.6 3,41996-2003 annual 26 9 1.9 2.4 2.4 1.7 9.2 5July 2002 - June 2003 30 38 6.4 8.0 8.5 - 28 6

ErieAug 1996 and May 1997 63 48 20 21 21 40 142 39 1,111996-2003 annual 84 31 10.0 11 7.7 7.9 26 5July 2002 - June 2003 24 22 8.7 9.4 8.4 - 50 61996-1998 annual ΣHCH 82 ΣChlordane 38 19 8

OntarioSummers of 1998 and 2000 48 37 6.9 8.2 6.2 5.0 37 22 1,11Jan - Dec 1992 4.9 12Oct 1995 - Sep 1997 3.8 12May 2000 - May 2001 48 35 6.9 9.3 4.0 8.5 12 3,41996 - 2003 annual 28 12 3.6 3.9 4.0 2.9 13 5July 2002 - June 2003 22 32 20 19 15 - 47 6July 2000 - June 2001 65 83 39 32 25 - 51 13,14

Michigan1996-1998 annual ΣHCH 104 ΣChlordane 14 15 81996-2003 annual 78 47 6.1 6.5 5.3 10 24 51996-2003 annual, Chicago 52 35 46 39 23 28 110 5July 2002 - June 2003 29 26 41 38 21 - 110 62002-2003 ΣChlordane 39.0 23 15summers of 1997 and 1998 13 10Sep-98 34 10

Southern Ontario May 2000-May 2001 58 81 4.4 3.1 0.87 1.1 4.0 3,4(inland) July 2000 - June 2001 55 165 17 17 17 - 72 13,14

July 2002 - June 2003 25 33 5.7 6.5 6.3 - 54 6

ArcticResolute Bay Summer 1992 114 9.8 0.51 1.4 0.77 - - 6.9 16Resolute Bay Summer 1999 44 7.4 1.4 1.4 0.80 0.70 2.0 19 11Canadian Archipelago Summer 1999 42 11 0.78 1.01 0.60 3.5 - 20 17,11

Bering/Chukchi Sea 1993 summer 91 23 0.76 1.2 0.60 1.43 1.1 4.1 18,11Bering/Chukchi Sea 1994 summer 125 18 0.92 1.9 0.81 - - 18,11Central Arctic Ocean 1994 Summer 64 14 0.44 0.67 0.46 - - 18,11North Atlantic 1994 summer 131 15 0.90 0.43 0.85 - - 18,11

Alert Jan-Jun 1992 1.6-27 19Alert 1993 Annual 58 10 0.4 1.0 - - 1.2 11 20Alert 1996-1998 ΣHCH 37 ΣChlordane 2.0 - - - - 21Alert 1998 2.7-5.1 22Alert 1993-2004 22-57 1.3-13 - - - - - - 23,24

Zeppelin 1994-2005 16-63 2.4-16 - - - - - - 25Storhofoi 1996-1998 ΣHCH 22 ΣChlordane 0.5 - - - - 21Ny-Alseund 1996-1998 ΣHCH 56 ΣChlordane 1.9 - - - - 21Bear Island 2000-2003 14 5.4 - - - - - - 26Bear Island 2000-2001 - - 0.96 1.1 1.1 - - - 26


1: Paper III; 2: Paper IV; 3: Shen et al., 2004; 4: Shen et al., 2005; 5: Sun et al., 2006; 6: Gouin et al., 2005; 7: Perlinger et al., 2005; 8: Buehler et al., 2001; 9: Glassmeyer et al., 1999; 10: James et al., 2001; 11: Jantunen, unpublished data; 12: Shoeib et al., 1999; 13: Harner et al., 2004; 14: Motelay-Massei et al., 2005; 15:Hoh and Hites, 2004; 16: Bidleman et al., 1995; 17: Paper VI; 18: Paper V; 19: Barrie et al., 1993; 20: Hargrave et al., 1997; 21: AMAP, 2004; 22: Braekevelt et al, 2001;23: Hung et al., 2002; 24: Halsall et al., 1998; 25: Becker et al., 2008; 26: Kallenborn et al., 2007.

36

Table 5: Dissolved Organochlorine Pesticides in Water (pg L-1)

Location Year αααα-HCH γγγγ-HCH TC CC TN HEPX Dieldrin Toxaphene Ref

Great LakesLake Superior 1996-1997 2760 612 4.1 5.4 3.2 45 137 918 1,2Lake Superior 1996 - - - - - - - 1120 3,4Lake Superior 1997-1998 - - - - - - - 910 5,6Lake Superior 1997 - - - - - - - 801 7Lake Superior 1998 - - - - - - - 696 8Lake Superior 1997-1998 2040 439 20 11 - - 145 - 9Lake Superior 2001-2005 1500 310 7.0 6.0 - - 110 - 10Lake Superior 2002 - - - - - - - 718 11Lake Superior 2005 1546 422 2.3 3.5 2.4 51 116 1041 12

Lake Michigan 1997 398 119 4.7 7.1 - - - - 9Lake Michigan 1997-1998 372 13

Lake Huron 1994-1995 - - - - - - - 250 3,4Lake Huron 1996-1997 1140 943 - - - - - - 1Lake Huron 1997-1998 465 92 1.4 3.4 9Lake Huron 2001-2005 170 130 2 3 - - 83 - 10Lake Huron 2005 235 161 - - - - - - 12

Lake Erie 1993 - - - - - - - 230 3,4Lake Erie 1996 480 491 - - - - - - 1Lake Erie 1997-1998 322 252 9.3 12 143 9Lake Erie 2001-2005 140 170 5.0 6.0 - - 120 - 10

Lake Ontario 1993 806 357 - - - - - - 13Lake Ontario 1993 170 3,4Lake Ontario 1998-2000 309 261 2.3 3.6 1.8 15.3 84 81 1,12Lake Ontario 2001-2005 190 180 nd nd - - 150 - 10-Fort Erie and 1986-1997 ΣChlordane <nd-140 - - 350-150 - 14Niagara on the Lake

ArcticBering Sea 1990 1500 190 1.5 1.9 0.5 - - - 15Bering Sea 1993 2002 454 1.0 0.8 0.5 2.5 3.7 20 16, 17Chukchi Sea 1993-1994 2060 430 1.7 1.2 0.76 6.3 - 28 16Chukchi Sea 2000 ΣHCH 2106 ΣChlordane 10 188 18Chukchi Sea 1997-1998 ΣHCH 1000-2000 ΣChlordane 2-18 20-22 19Chukchi Sea 1998 40 20

Resolute Bay 1992 4700 440 7.3 4.5 1.5 - - 48 21Resolute Bay 1992 3640 520 0.5 1.3 - - 12 85 22Amituk Lake 1992 1300 280 3.3 4.1 1.9 - - 145 21Beaufort Sea 1992-1993 3500-5500 500-700 20Canadian Archipelago 1999 3500 310 1.4 1.5 1.1 1.8 15 113 23,12Canadian Archipelago 1999 ΣHCH 2277 ΣChlordane 8 150 18Baffin Bay 1998 ΣHCH 1233 ΣChlordane 15 253 18Hudson Bay 1999 ΣHCH 1687 ΣChlordane 11 24

Eastern Arctic Ocean 1996 25,26Central Arctic Ocean 1994 2420 470 1.3 1.3 0.8 14.8 - 79 16Western Arctic Ocean 1992-1993 2000-3000 500-1100 27,28Greenland Sea 1994 870 200 1.6 0.84 1 6.6 - 126 16

White Sea 1999-2000 41 29


1: Paper III; 2: Paper IV; 3: Swackhamer et al., 1998; 4: Swackhamer et al., 1999; 5: James et al., 2001; 6: Swackhamer and Symonik, 2004; 7: Burniston and Strachan, 2001; 8: Muir et al., 2004; 9: Buehler et al., 2000; 10: Blanchard et al., 2008; 11: Muir et al., 2003; 12: Jantunen et al., unpublished; 13: Ridal et al.,1996; 14: Niagara River Secretariat, 2000; 15: Iwata et al.,1993; 16: Paper VII; 17: Strachan, unpublished; 18: Hoekstra et al.,2002; 19: AMAP,2004; 20: Macdonald et al., 2001; 21: Bidleman et al., 1995; 22: Hargrave et al., 1998; 23: Bidleman et al., 2007; 24: Hoekstra et al., 2003; 25: Harner et al., 1999; 26: Strachan et al., 2000; 27: Macdonald et al., 1996; 28: Jensen et al., 1997; 29: Muir et al., 2002.

37

Non racemic residues of chlordanes were found in the Great Lakes air and water, with the same pattern of

(+)trans-chlordane and (–)cis-chlordane depletion as seen in agricultural soils of the Great Lakes Basin (Bidleman et

al., 1998a; Gouin et al., 2007; Ulrich and Hites, 1998; Shen and Wania, 2005b; Paper III). When air samples were

taken at multiple sites within the Great Lakes Basin, a trend was noted for EFs of trans-chlordane and cis-chlordane to

be closer to racemic values in urban areas and non-racemic in rural areas (Gouin et al., 2007; Shen and Wania,

2005b). Trans-chlordane/cis-chlordane ratios were positively correlated with the EF of trans-chlordane and

negatively correlated with the EFs of cis-chlordane (Gouin et al., 2007). These “fresher” chlordane signatures reflect

the tendency of chlordanes in urban air to be less influenced by microbial degradation processes, possibly because

they are emitted more from former termiticide usage sites than from soils. Transport of chlordanes with different EF

signatures to the lakes can be expected from different sources; however, a lake vs. terrestrial source apportionment

based on EFs, as for α-HCH, is difficult because the EFs of chlordanes in background air are more variable and

generally nonracemic.

6.3. Arctic

Toxaphene is of concern in the Arctic because it is typically the most abundant OCPs in biota. Fish (Evans et

al., 2005), beluga (Andersen et al., 2006; Braune et al., 2005; Stern et al., 2005), narwhal (Dietz et al., 2004 ) and

terrestrial plants (France et al., 1997) are higher in Σtoxaphene than other OCPs such as DDTs, chlordanes and HCHs.

Ringed seal tend to have lower levels of Σtoxaphene compared to DDTs (Cleemann et al., 2000). Toxaphene has also

been found in plankton and amphipods (Braune et al., 2005; Hoekstra et al., 2002;) and terrestrial animals (Hoekstra

et al., 2003a,b).

Cyclodienes are bioaccumulative and have been found in all types of arctic biota including invertebrates,

freshwater and marine fish, terrestrial carnivores, seabirds and marine mammals (AMAP, 2004; Braune et al., 2005).

Circumpolar trends of chlordanes in ringed seals show higher concentrations in east Greenland, the northern Canadian

Archipelago, Beaufort Sea and the Barents Sea while lower concentrations occur in Hudson Bay, the southern

Archipelago and the White Sea (AMAP, 2004).

Like toxaphene and cyclodienes, HCHs bioaccumulate in arctic biota and have been found in invertebrates,

marine fish, seabirds, ring seals, beluga, narwhal and polar bears (Braune et al., 2005). Fisk et al. (2001) found that

ΣHCHs bioaccumulate but to a lesser extent than other OCPs and bioaccumulation was mainly of the β-isomer.

ΣHCH concentrations in ringed seal from Ikpiarjuk (Arctic Bay), Canada decreased from 1975-2000, but the

proportion of β-HCH/ΣHCHs increased (Braune et al., 2005).

6.3.1. Air

Toxaphene has been found in all abiotic compartments of the Arctic, air, seawater, suspended sediments and

snow (Macdonald et al., 2000). Very few measurements of Σtoxaphene have been made in arctic air, see Table 4 for

summary. These show higher concentrations in the late 1980s compared to the 1990s (Li and Macdonald, 2005, see

Figure 10 in their paper). In the 1990s (Table 4) concentrations ranged from 1.6-20 pg m-3, where the higher

38

concentrations were seen in the summer of 1999 on TNW’99 (Barrie et al., Bidleman et al., 1995; Hargrave et al.,

1997; Jantunen et al., unpublished data). Braekevelt et al. (2001) determined the Σtoxaphene in arctic air at Alert NT,

which ranged from 2.7-5.1 pg m-3 depending upon the standard used. They also examined the homologue distribution

in air and found that air samples were enriched in the heptachloro- homologues relative to technical toxaphene; this

was also as seen in Alabama and the Great Lakes air samples (Paper II; Paper IV).

The concentration of toxaphene in arctic air cannot be reconciled by only looking at North American usage

(MacLeod et al., 2002) and is probably more controlled by northern hemisphere cycling processes and usage (past and

current). Toxaphene amounts comparable to those in North America were used in Eastern Europe and the former

Soviet Union, and these sources must also be considered as contributing to the arctic contamination. The closest

North American usage of toxaphene to the arctic was to eradicate rough fish in Yukon lakes, but amounts applied

were very small compared to total usage (Donald et al., 1998; Miskimmin et al., 1994; 1995). Possible sources to the

Arctic are atmospheric transport from countries where toxaphene is currently used or volatilization from past usage

regions. The former seems unlikely today, since toxaphene is undergoing worldwide elimination under the Stockholm

Convention. However, contaminated soils remain a significant source of toxaphene emissions (Bidleman and Leone,

2004a,b; Li et al., 2001; MacLeod et al., 2002).

Air monitoring has been ongoing in the Arctic under the Arctic Monitoring and Assessment Programme

(AMAP) since the late 1980s-early 1990s. During 2000-2003, six monitoring stations collected weekly samples for

compounds including dieldrin, chlordanes and its metabolites (Su et al., 2008). Concentrations of these cyclodienes

in arctic air were lower than in the Great Lakes region (Paper III) and southern U.S.A. (Paper II), see Table 4.

Dieldrin in arctic air showed strong seasonal and spatial variations and significant temperature correlations were

found for about half of the arctic stations for the cyclodiene compounds (Su et al., 2008). Levels of the Σchlordane at

monitoring stations were similar, indicating that arctic air is well-mixed and current sources are from secondary

volatilization emissions not primary current use emissions (AMAP, 2004). This is supported by proportions of trans-

chlordane and cis-chlordane enantiomers in arctic air which are closer to non-racemic residues in soils than racemic

chlordanes in the technical product and termiticide-related sources (Bidleman et al., 2002; 2004). Chlordane-related

compounds and dieldrin are declining at Alert (82.5N, 62.2W) in the Canadian Arctic, half lives range from 5.5-8.3

and 13 years, respectively (Hung et al., 2005) and slight decline were also observed at Svalbard and Ny Ålesund

between 1994-2005 (Becker et al., 2009).

Cyclodienes are scavenged from the air by precipitation and have been found in arctic snow and ice from

Spitsbergen, Greenland, Russia and Canada (Herbert et al., 2006, Hermanson et al., 2005; Macdonald et al., 2000)

although they are not consistently found and vary widely in concentration (Herbert et al., 2006). Heptachlor exo-

epoxide and dieldrin were found in ice cores from Spitsbergen, where maximum concentrations were found between

1992-1998 and 1979-1986 respectively (Hermanson et al., 2005) and had higher concentrations than in the Agassiz

ice cap in the Canadian Arctic (Gregor et al., 1989).

Compared to other organic contaminants, HCHs have a high long range transport potential (Leip and Lammel,

2004) and the arctic atmosphere quickly adjusts to changes in global emissions. This has been shown by the strong

relationship between drops in global usage/emissions of HCHs and declining levels in arctic air (Li and Bidleman, 2003).

Two stepwise drops in air concentration correlate with major reductions in usage, the first was in 1983 when China

39

banned the usage/production of technical HCH and the second was in 1990 when the former U.S.S.R. stopped using

technical HCH and India stopped its use on food crops, (Li, 1999a; Li 1999b; Li et al., 2000).

Atmospheric measurements in the late 1990s-2003 showed fairly uniform concentrations of α- and γ-HCH across

the Arctic with some regional differences (Table 4) and γ-HCH was generally 15-20% of α-HCH (AMAP, 2004; Shen et

al., 2004; Su et al., 2006). Half lives for α-HCH in arctic air in the 1990s ranged from 9-17 years, where γ-HCH had a

shorter half life of 4.9-5.7 years (Hung et al., 2005). In Paper III, a three-fold decrease was seen for HCHs in air samples

taken in 1988 and 1993-1994 in the Bering and Chukchi Seas. During AOS-94, α-HCH concentrations above the open

water of regional seas at lower latitudes were higher than at more northern latitudes over ice-covered waters. In Paper VI,

levels of α-HCH averaged 45 pg m-3 in the Canadian Archipelago during spring-summer 1999, but concentrations at

Resolute Bay were lower in June-July and higher in late July and August. This was due to the ice breaking up and release

α-HCH from the water (see below). In contrast, γ-HCH concentrations in Archipelago air did not vary significantly over

spring-summer.

6.3.2. Water

There have been few studies of toxaphene in arctic seawater. In Paper VII, Σtoxaphene was measured in

surface seawater on a 1994 transect across the central Arctic Ocean and into the Greenland Sea. Concentrations were

higher in the central Arctic Ocean and Greenland Sea compared to the Chukchi Sea (Paper VII). Literature data in

Table 5 indicates that Σtoxaphene concentrations are higher in Baffin Bay > Beaufort Sea ~ Canadian Archipelago >

White, Bering and Chukchi seas. This is consistent with air pathways from the southern U.S., which travel northeast

and sweep over the Great Lakes and eastern Canada (Ma et al., 2005a,b).

In Paper VII, toxaphene congeners B8-1413 and B9-1679 were quantitatively determined in water samples

from the western Arctic Ocean. Qualitatively, chromatograms showed that the labile octachlorobornanes B8-531 and

B8-806/808 appeared depleted in water and that the stable congener B8-1413 was enriched in water compared to the

technical toxaphene mixture. Similar enrichments and depletions were found in 1993 air samples from the Bering and

Chukchi seas (Jantunen, 1997). Different toxaphene congener patterns were seen in water samples taken from the

White Sea compared to the southern Beaufort Sea (AMAP, 2004). Beaufort water had higher levels of Σtoxaphene

but showed a more highly degraded-older pattern dominated by terminal residues, whereas the White Sea had lower

concentrations but more highly chlorinated congeners. The pattern in the White Sea more resembles the pattern seen

in Chukchi sea and central Arctic water (Paper VII), showing depletions of the labile congeners discussed above.

Water samples from Holman, Northwest Territories and Barrow Alaska were dominated by B6-923, a terminal

congener (Hoekstra et al., 2002) and showed a different congener distribution, not depleted in the labile congeners

from above. The former U.S.S.R. used a polychloroterpene mixture, so the differences in pattern may be due to a

different congener distributions in the Russian and U.S. toxaphene mixtures.

Very few reports of cyclodienes in Arctic Ocean water are available, especially in recent years, see Table 5 for

summary. ΣChlordanes in Arctic Ocean water were in the same range as in the Great Lakes, 2-18 pg L-1, whereas

dieldrin and heptachlor exo-epoxide concentrations were lower ranging from 3-22 pg L-1. In Paper VII concentrations

of trans-chlordane, cis-chlordane and heptachlor exo-epoxide were measured in Arctic Ocean surface water on AOS-

40

94. Heptachlor exo-epoxide increased from the Chukchi Sea to the pole and then decreased toward Spitsbergen and

Greenland Sea. Little variation with latitude was seen in chlordane concentrations, although the ratio of trans-/cis-

chlordane decreased at high latitudes.

At Resolute Bay, NU, Canada, (74.5N, 94.9W) water samples were collected in the summer of 1992 by

Bidleman et al. (1995) and over a year in 1993-1994 by Hargrave et al. (1997). Concentrations of trans-chlordane,

cis-chlordane and dieldrin were slightly higher in winter-early spring compared to summer-fall. The trans-

chlordane/cis-chlordane ratio ranged from 0.4-1.6 with no clear trend (Table 5).

Macdonald et al. (2001) measured vertical distributions of chlordanes and dieldrin in the water column of the

Canada Basin and Chukchi Plateau. In February and September 1998, the depth profiles were noisy with no trend

with depth from the surface to 250m, whereas in April and July concentrations were lower at the surface and increased

with depth (AMAP, 2004). Different profiles, with decreasing levels with depth over 275 m, were found also by

Hargrave et al. (1988) in the Beaufort Sea.

In Paper VII the enantiomers of trans-chlordane, cis-chlordane and heptachlor exo-epoxide were reported.

These compounds were racemic in Arctic Ocean water when sampled in 1994. This might be due to transport and

deposition of fresh chlordanes in the Arctic during the years of heavier chlordane use, where due to low temperatures

and low microbial activity the chlordanes have remained largely undegraded. This was also seen by in water samples

reported in Hoekstra et al. (2003). Heptachlor undergoes degradation in soils to heptachlor exo-epoxide, with

preferential formation of the (+) enantiomer, and subsequent volatilization leads to enrichment of the (+) enantiomer

in ambient air (Bidleman et al., 1998b). Atmospheric transport delivers non-racemic heptachlor exo-epoxide to the

Arctic where it is atmospherically deposited. This accounts for the observed enrichment of (+) heptachlor exo-

epoxide in Arctic Ocean water (Paper VII).

HCHs are the most abundant OCPs in Arctic Ocean water. (CACAR, 2003, AMAP, 2004). Mass budgets of

HCHs in the arctic have been reported by Barrie et al. (1992), Wania et al., (1999a,b), Macdonald et al. (2000), Li et al.

(2002, 2004) and Toose et al. (2004). Historically, the largest α-HCH loadings to the arctic was from atmospheric

deposition followed closely by ocean currents and rivers but currently ocean transport is dominant (Li et al., 2004). For

β-HCH, ocean transport has always been the dominant pathway (Weber et al., ; Li et al., 2004). Peak loadings and

concentrations of α-HCH in the Arctic Ocean occurred in 1982 (Li, 2004; Wania et al., 1999). Currently the Arctic

Ocean is experiencing a net loss of HCHs, where the largest removal mechanism is microbial degradation followed by

ocean currents and transport to deep water (Macdonald et al., 2000; Li et al., 2004).

Large spatial variations exist for HCHs in Arctic Ocean surface waters (Table 5). Lowest levels of α-HCH occur

in the eastern Arctic Ocean and its subarctic seas (AMAP, 2004; Harner et al., 1999; Lakaschus et al., 2002; Paper V,

VII) and the Greenland Sea (Paper V, VII), while the highest levels are in the Beaufort Sea and western Canadian

Archipelago (Bidleman et al., 2007). Intermediate concentrations occur in the Bering/Chukchi seas and northern Canada

Basin (Paper V and VII, AMAP, 2004; Hoekstra et al., 2002; Iwata et al., 1993; Macdonald et al., 2001). These

concentration gradients were created and maintained by complex circulation in the Arctic Ocean and lack of horizontal

and vertical mixing. Pacific water entering the Bering Strait was impacted by atmospheric deposition and ocean transport

from high usage areas in Asia during years when the emissions were highest. This water, containing the highest

concentrations of α-HCH, is currently circulating around the Beaufort Gyre under a cap of ice. The ice cap reduces

41

removal by volatilization. The main outflow pathways of HCHs from the Arctic Ocean is percolation of Beaufort Sea

water through the Canadian Archipelago causing a decreasing west to east concentration gradient (Bidleman et al., 2007;

Paper VI). Outflow also takes place via the East Greenland Current (Li et al., 2004; Macdonald et al., 2000).

Microbial degradation in the water column, one of the largest loss mechanisms, can be traced by determining

the enantiomers of α-HCH. Preferential loss of (+)α-HCH increases with depth in the Arctic Ocean (Paper V and

VII, Harner et al., 1999). This was first noted by Jantunen and Bidleman (1996; Paper V), who collected samples at

different depths in the central Arctic Ocean and regional seas in 1993-1994. An example of depth profiles of EFs is

shown in Figure 11 for four arctic-subarctic stations. Figure 11 shows profiles in different Arctic locations. Profiles

of α-HCH are influenced by the structure of the water column. The Greenland Sea is in a gyre that mixes the water

column, shown by the small changes in EFs with depth. Where water in the Canada Basin is stratified, the upper most

layer is a cold low salinity mixed polar layer, below that is broad halocline. The deeper thus older water masses show

highly degraded α-HCH patterns. Harner et al. (1999) used profiles of decreasing concentrations and EFs with depth

in the Barents Sea and eastern Arctic Ocean to calculate the pseudo first-order microbial degradation rate constants of

α-HCH enantiomers. Estimated half-lives due to microbial degradation were 5.9, 23 and 19 y for (+)α-HCH, (–)α-

HCH and γ-HCH, respectively. These are shorter than half-lives due to base hydrolysis in seawater (Harner et al.,

2000). In arctic lakes, enantioselective and non-enantioselective microbial degradation are important loss processes.

Helm et al. (2000) reported that the (+) enantiomer was preferentially degraded in three streams in the Canadian

Arctic. Most enantioselective degradation of α-HCH occurred during peak runoff in the streams but EFs were lowest

in the lakes in late summer. A mass balance indicated that 33-61% of the α-HCH was lost within the lake by microbial

degradation that was not enantioselective, leading to a half life ranging from 0.61-1.44 y.

6.3.3. Air-Water Gas Exchange

With the decrease of HCHs in arctic air, discussed in Section 6.3.1, the water/air fugacity ratios reversed between

1988 and 1993 from net deposition to volatilization in the Bering and Chukchi Seas for α-HCH and from net deposition

to near equilibrium for γ-HCH (Jantunen and Bidleman, 1995). During the 1994 Arctic Oceans Sections (AOS-94)

expedition, air and water samples were collected in subarctic and arctic regions to determine the air-water gas

exchange direction and magnitude (Paper V, see Figure 1 for the cruise track).

Higher concentrations of α-HCH in air were found over areas of open water in the Bering, Chukchi and

Greenland seas, while lower concentrations were found over the ice-covered Central Arctic Ocean. Over open water,

enantiomer profiles of α-HCH (given as ERs in Paper V) in air matched those in surface water and reflected the

different enantioselective degradation among regions, i.e. depletion of (+)α-HCH in the eastern Arctic Ocean but

depletion of (–) in the Bering-Chukchi seas. The EFs in air over ice-covered portions of the Arctic Ocean remained

nearly racemic, despite nonracemic α-HCH in seawater under the ice. The ice pack caused the air and water to be

decoupled due to the lack of air-water gas exchange (Figure 12, where ERs from Paper V have been converted to

EFs). This was again observed at Resolute Bay in 1999. In Paper VI, fugacity ratios were calculated from TNW-99

parallel air and water samples and at Resolute Bay where air samples were paired with water samples collected in the

central Archipelago. At Resolute Bay, α-HCH was near equilibrium or undergoing net volatilization from the surface

42

0

0.1

0.2

0.3

0.4

0.5

0.6

0 100 200 300 400 500 600 700

Depth (m)

EF

Chukchi Sea Canada Basin North Pole Greenland Sea

Figure 11: EF of α-HCH in water with depth in the Arctic Ocean at four stations. Showing differing profiles in different locations. Profiles are influenced by the structure of the water column. The Greenland Sea is in a gyre that mixes the water column, where in the Canada Basin the water column is very structured into layers, where older water masses show highly degraded α-HCH patterns.

water, while γ-HCH was near equilibrium or undergoing net deposition. Until mid-July the net flux at Resolute Bay was

zero because the surrounding ocean was ice covered. When the ice broke up, α-HCH was released into the air, increasing

the concentration by ~30%. In addition to the air concentration increase the EF of α-HCH in the air changed, decreasing

from close to the racemic value of 0.496 before ice breakup to 0.483. The EF of α-HCH in surface water in the central

Archipelago was 0.444. Application of eq 21 indicated that 32% of α-HCH in air after ice breakup was due to

volatilization. In contrast to α-HCH, γ-HCH did not show an increase after ice break up, it showed a slight decreased,

fugacity ratios predict that γ-HCH is being deposited into the surface water.

The EFs of α-HCH in different Archipelago water masses along the TNW-99 cruise track ranged from 0.432-

0.463. The EFs of α-HCH in air samples taken from the ship over open-water areas were closely correlated to those in

water (r2 = 0.68), while EFs over ice-covered areas were higher and were not correlated to those in water.

In other studies, Ding et al., (2007) found that α- and γ-HCH were near equilibrium or volatilizing from the

North Pacific Ocean, Bering Sea and Chukchi Sea. Fugacity ratios of samples taken in the eastern Arctic Ocean from

1994-2001, predicted near-equilibrium for α-HCH (Harner et al. 1999; Lakaschus et al., 2002) but showed a varied

picture for γ-HCH, one study indicated net deposition (Harner et al., 1999) while another indicated near equilibrium

(Lakaschus et al., 2002).

43

Figure 12: EF of α-HCH in air from AOS-94.

Hargrave et al. (1997) collected water samples at Resolute Bay and air samples at Alert, and found a seasonal

cycling in the net air-water gas exchange direction of OCPs. This was due to relatively large annual cycle of OCPs

in air and to a lesser extent, to smaller seasonal variations in seawater concentrations. FRs for α-HCH and dieldrin

were always >1 (net volatilization). FRs for γ-HCH were also >1 in most months, but dipped below 1 (net

deposition) in May-April, coinciding with a springtime peak in air concentrations and usage in North American

temperate regions. Net deposition of toxaphene was noted in spring-summer, in response to higher air

concentrations in these seasons, and returned to net volatilization in fall-winter. Similar trends were found for the

chlordanes. Although these FRs indicate the potential for gas exchange, flux calculations were meaningful only in

the summer when no ice cover was present.

References Agency for Toxic Substances and Disease Registry (ATSDR). 2002. Toxicological Profile for Aldrin/Dieldrin. U.S. Department of Human and Health Services, Public Health Service, Atlanta, GA. Agency for Toxic Substances and Disease Registry (ATSDR). 1996. Toxicological Profile for Toxaphene. U.S. Department of Human and Health Services, Public Health Service, Atlanta, GA. Agency for Toxic Substances and Disease Registry (ATSDR). 1994. Toxicological Profile for Chlordane. U.S. Department of Human and Health Services, Public Health Service, Atlanta, GA.

Ice CoverOpen Water Open Water

Bering/Chukchi Seas Central Arctic Ocean Greenland Se a

0.40

0.42

0.44

0.46

0.48

0.50

0.52

0.54

0.56

50-5

455

-59

60-6

465

-69

70-7

475

-79

80-8

485

-89

89-8

584

-80

77-6

767

-61

55-4

5

Latitude (N)

EF EF Air

EF Water

Ice CoverOpen Water Open Water

Bering/Chukchi Seas Central Arctic Ocean Greenland Se a

0.40

0.42

0.44

0.46

0.48

0.50

0.52

0.54

0.56

50-5

455

-59

60-6

465

-69

70-7

475

-79

80-8

485

-89

89-8

584

-80

77-6

767

-61

55-4

5

Latitude (N)

EF EF Air

EF Water

44

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Paper 1

Henry’s Law Constants for Hexachlorobenzene, p,p’-DDE and Components of Technical Chlordane and

Estimates of Gas Exchange for Lake Ontario

Liisa M. Jantunen and Terry F. Bidleman

Chemosphere, 2006, 62, 1689-1696. Centre for Atmospheric Research Experiments (CARE), Environment Canada 6248 Eighth Line, Egbert, Ontario, Canada, L0L 1N0. Contributions: Liisa Jantunen did the experimental determination of Henry’s Law constants, prepared, collected and analysed the Lake Ontario water and air samples. Liisa carried out the data analysis and wrote the paper. Terry Bidleman secured funding and provided scientific guidance during every step of this project.

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Abstract

The Henry's Law constants (HLC) for trans- and cis-chlordane (TC, CC), trans-nonachlor (TN),

hexachlorobenzene (HCB) and p,p’-DDE were determined by the gas stripping method over a temperature range

of 5 to 35 oC. The HLC variation versus temperature (K) was described by log H = m/T + b. Parameters of this

equation were (with standard deviations) TC: m = -1524±158, b = 6.58±0.54; CC: m = -1786±209, b =

7.42±0.71; ΤΝ: m = -2068±284, b = 8.44±0.97; HCB: m = -3013±174, b = 11.60±0.59 and p,p’-DDE: m = -

2043±240, b = 8.37±0.82. The HLCs (Pa m3 mol-1) at 25oC (298.15 K) were: TC = 29; CC = 27; TN = 32; p,p’-

DDE = 33 and HCB = 35. These HLCs values were used to calculate fugacity ratios from paired air and water

data from Lake Ontario, July 1998. The resulting fugacity ratios predict that volatilization was occurring for all

compounds during that month.

1. Introduction

Chlordanes, p,p’-DDE and hexachlorobenzene (HCB) are abundant organochlorine compounds (OCs) in the

North American Great Lakes and arctic ecosystems and the Henry's Law constant (HLC, Pa m3 mol-1) is a critical

property for describing their partitioning between air and water. The HLC of a compound (H) is the ratio of its

partial pressure in air (P, Pa) to its concentration in water (Cw, mol m-3) at equilibrium:

wC

PH = (1)

For compounds that are slightly soluble, the HLC can be calculated from the ratio of the saturation vapour

pressure (Psat) and water solubility (Cwsat)

satW

sat

= C

PH (2)

However, it is often the case that literature values for these properties are quite variable and lead to large

computational differences in predicted H values, so it is preferable to measure the HLC directly. Several

different methods are employed to determine HLCs: wetted wall column (Altschuh et al., 1999; Fendinger and

Glotfelty, 1988, 1989; Gautier et al., 2003; Shepson et al., 1996; Rice et al., 1997), static headspace sampling

(Murphy et al., 1987), fog chamber (Fendinger and Glotfelty, 1989) and gas stripping. Two versions of the latter

technique are common. 1. Conventional gas stripping: the chemical is purged from a column of water with air or

nitrogen and the decrease in water concentration is followed over time (Alaee et al., 1997; Breitner et al., 1998;

Kucklick et al., 1991; Mackay et al., 1979; Sahsuvar et al., 2003; ten Hulscher et al., 1992). 2. Dynamic

headspace sampling: the chemical is measured concurrently in the water and gas stream exiting the system

(Bamford et al., 1999; 2000; Sahsuvar et al., 2003; Yin and Hassett, 1986). The latter version is advantageous

for chemicals which strip slowly due to low HLC values.

Previous reports of directly measured HLCs for p,p’-DDE and the technical chlordane components, trans-

chlordane (TC), cis-chlordane (CC) and trans-nonachlor (TN) have been at a single temperature (Atlas et al.,

63

1982; Fendinger and Glotfelty, 1989; Warner et al., 1987). The variation in the HLC of HCB has been reported

as a function of temperature by only one laboratory (ten Hulscher et al., 1992). In this work the HLCs of these

compounds were determined over a range of 5-35 oC.

2. Materials and methods

2.1. Gas stripping experiments

The HLCs were determined using the Version 1 of the gas-stripping method. The apparatus was the same as

that described and used by Kucklick et al. (1991) and subsequently used by Jantunen and Bidleman (2000) and

Sahsuvar et al. (2003). The apparatus, shown by Kucklick et al. (1991) in Figure 1 of that paper, is a three-

chamber annular vessel. The outside chamber contains air for insulation, the middle chamber contains water

circulated from a constant temperature water bath at a controlled temperature and the innermost chamber

contains the aqueous solution of the pesticides where the gas stripping takes place. The volume and height of the

inner chamber are 525 ml and 62 cm. The purge tube, with a coarse frit on the end, is lowered to within a few

millimeters of the bottom of the inner chamber, 47 cm below the initial water surface. Kucklick et al. (1991)

showed that the air bubbles achieved equilibrium with hexachlorocyclohexanes (HCHs) in water by sampling at

two different depths (45-50 and 26 cm) and found no significant differences in the HLCs. The top of the inner

chamber narrows and coils three times, then leads to the outlet. The coil is to prevent aerosols from escaping the

system. Jantunen and Bidleman (2000) carried out an experiment to check this possibility. The water in the inner

chamber was spiked with fluorescein, a glass wool trap was installed on the outlet and the apparatus was allowed

to bubble overnight at room temperature. A UV light was used to check for the presence of fluorescein on the

glass wool, but none was found. Temperatures are measured in the inner chamber and the water bath temperature

is adjusted to achieve the desired value.

In the present experiment, the inner chamber was filled with HPLC grade water that was spiked with 1000 µl

of acetone containing 10.4 - 21.9 ng µl-1 of TC, CC, TN, HCB, and p,p’-DDE (U.S. Environmental Protection

Agency Repository for Pesticides and Industrial Chemicals, Research Triangle Park, NC, U.S.A.). This gave

initial concentrations in water of 19.8 - 41.7 ng ml-1. These concentrations were below reported liquid-phase

water solubilities for p,p'-DDE (250 ng ml-1), CC (530 ng ml-1) and TC (612 ng ml-1) (Mackay et al., 1999) and

for HCB (570 ng ml-1) (Final Adjusted Values, Shen and Wania, 2005). Zero grade ultra high purity nitrogen at

a flow rate of 0.02 - 0.025 m3 h-1 was introduced through a coarse glass frit at the bottom of the inner chamber to

purge the pesticide vapours. The flow rate was taken frequently throughout the experiments by a bubble flow

meter and typically varied by 2-3 % over an experiment. Evaporation loss of water was minimized by passing

the nitrogen through a bubbler containing HPLC grade water and placed in the constant temperature bath to

saturate the gas stream before entering the stripping chamber. The system was allowed to equilibrate at the

experimental temperature and with the nitrogen flowing for 15 - 30 min before the first sample was taken. The

stripping experiments were carried out at 5, 10, 20, 30 and 35 oC, and the water was purged for 4 - 10 h

depending on the temperature; shorter purge times were needed at higher temperatures because the compounds

were more quickly depleted. Water samples were collected by withdrawing two aliquots of 5.0 ml from the

64

bottom outlet; the first aliquot (discarded) was to flush the stopcock and stem and the second was kept for

analysis.

2.2. Analysis

Extraction of the 5.0 ml water samples was done by adding 1.0 ml of isooctane, and 10 µl of 9.7 ng µl-1 mirex

internal standard. The sample was vortex mixed for 1 min, centrifuged and the top layer of isooctane was

removed. The water was extracted with another 1 ml of isooctane and the two extracts were combined. The

isooctane extraction efficiency was checked by extracting all water samples from the first experiment (n=12) with

an extra 1 ml portion of isooctane; the concentrations of analytes in this extract were below the method detection

limit. Extracts were analysed by gas chromatography using a Hewlett Packard 5890 GC with a 63Ni electron

capture detector. Data collection and quantification were done with the HP Chemstation. The column was a DB-

5 (J&W Scientific Co., Rancho Cordova, CA, U.S.A., 60 m x 0.25 mm i.d., 0.25 µm film thickness). The

temperature program was: initial temperature 90 oC hold 1.0 min, 10 oC min-1 to 160 oC, 2.0 oC min-1 to 240 oC,

20 oC min-1 to 270 oC hold 10 min. A 2-µl splitless injection of the sample was made (split opened after 0.5

min.). Other conditions: injector temperature 250 oC, detector temperature 280 oC, H2 carrier gas linear velocity

40 cm s-1 and N2 detector makeup gas flow 1.1 ml s-1. Quantification was done against four standards (dilutions

of the original spiking solution, with internal standard) that spanned the range of sample concentrations.

2.3. Data treatment

The loss rate of the dissolved phase can be expressed by the integrated mass balance equation (Mackay et al.,

1979):

Ct and Co are the aqueous concentrations at time t and zero, G is the gas flow rate (m3 h-1), Vt is the solution

volume (m3) at time = t, T is the temperature (K) and R is the gas constant (8.314 Pa m3 mol-1 K-1). Water

losses due to evaporation ranged from zero to 20 ml and averaged 5.5 ml during 4-10 h purge times. Water

volumes removed from the system for sampling ranged from 80 to 120 ml.

When Vt log Ct /Co is plotted against t, the result is linear (Figure 1) with a slope of

tV

HG

C

C ] RT303.2

[ = log0

t − (3)

This equation was modified to account for variations in the water volume during the experiment, which were mainly

due to water samples being removed for analysis plus a small amount of evaporation (Breiter et al., 1998; Gossett et al.,

1987; Sahsuvar et al., 2003).

t

HG

C

CV

RT2.303 = log

o

tt

− (4)

65

–[HG/ 2.303 RT]. The temperature dependence of the HLCs was expressed by eq 5:

Log H = m/T + b (5)

Figure 1: Bubble stripping experiment at 10oC, for trans-chlordane (--▲--), cis-chlordane (--■--) and trans-

nonachlor(�●�).

Since the number of experiments at each temperature differed, plots were made of the average HLCs at each

temperature (Figure 2). The slope is related to the enthalpy of water-to-air transfer (∆Hwa, kJ mol-1) by:

m = –103∆Hwa/2.303R (6)

3. Results and discussion

3.1. Henry’s law constants

Results of the stripping experiments for TC, CC, TN, HCB and p,p’-DDE are given in Table 1 as the slope

(m) and intercept (b) of eq 5 and the HLC value at 25 oC. The HLCs at 25 oC of all five chemicals were quite

similar, ranging from 27 - 35 Pa m3 mol-1.

Comparison to literature values at 20-25oC is given in Table 2. Two other bubble stripping determinations

have been reported for the chlordanes. HLCs measured by Atlas et al. (1982) for TC and CC were 135 and 89 Pa

m3 mol-1 respectively, about 3 - 5 times higher than our values of 29 and 26 Pa m3 mol-1. Warner et al. (1987)

reported 4.9 Pa m3 mol-1 for "chlordane" (isomer unspecified). Fendinger and Glotfelty (1989) determined the

HLC of TC by wetted wall column and fog chamber techniques, with good agreement between the two methods

(HLCs 8.8 and 5.9 Pa m3 mol-1). Iwata et al. (1993) reported HLCs of 16.5, 11.3 and 49.2 Pa m3 mol-1 for TC,

CC and TN based on the ratio of vapour pressure to water solubility, however the individual vapour pressure and

-0.8

-0.7

-0.6

-0.5

-0.4

-0.3

-0.2

-0.1

0

0.00 1.00 2.00 3.00 4.00 5.00 6.00 7.00 8.00

t (h)

Vt Lo

g C

t/Co

66

solubility values were not given. Shen and Wania (2005) derived an internally consistent set of physical-

chemical property data for TC and CC. Their resulting Final Adjusted

Figure 2: Plots of Eq. (5) for HCB (a), trans-chlordane (b), trans-nonachlor (c) and p,p’-DDE (d).

Values (FAVs) for the HLCs of TC and CC were 6.8 and 5.7 Pa m3 mol-1, which are ~4 times lower than ours.

Few comparisons are available for p,p'-DDE. Our HLC of 35 Pa m3 mol-1 was higher than 4.2 Pa m3 mol-1,

determined by wetted wall column (Altschuh et al., 1999) but lower than 122 Pa m3 mol-1 reported from another

bubble stripping study (Atlas et al., 1982). Shen and Wania (2005) reported a FAV of 4.2 Pa m3 mol-1, which

they assumed to be the same as the Altschuh et al. result.

Several HLCs have been reported for HCB. Our result of 35 Pa m3 mol-1 agrees well with two other bubble

stripping results (41 Pa m3 mol-1, ten Hulscher et al., 1992; 49 Pa m3 mol-1, Oliver et al., 1985), a wetted wall

column determination (24 Pa m3 mol-1, Altschuh et al., 1999) and a result using equilibrium partitioning in a

closed system (26 Pa m3 mol-1, Hansen et al., 1993). Ours is much lower than two other bubble stripping results,

130 Pa m3 mol-1 (Atlas et al., 1982) and 170 Pa m3 mol-1 (Warner et al., 1987) and a factor of two lower than 71

Pa m3 mol-1, obtained by equilibrating water with a HCB-saturated airstream (Atlas et al., 1983). The FAV of

Shen and Wania (2005) was 65 Pa m3 mol-1.

R 2 = 0 .9 6 4

0.7

0.9

1 .1

1 .3

1 .5

1 .7

1 .9

2 .1

3.2 3.25 3.3 3 .35 3.4 3.45 3.5 3.55 3 .6 3.65

1000/T (K )

Lo

g H

R2 = 0.915

0.7

0.9

1.1

1.3

1.5

1.7

1.9

2.1

3.2 3.25 3.3 3.35 3.4 3.45 3.5 3.55 3.6 3.65

1000/T (K)

Lo

g H

R2 = 0.952

0.7

0.9

1.1

1.3

1.5

1.7

1.9

3.2 3.25 3.3 3.35 3.4 3.45 3.5 3.55 3.6 3.65

1000/T (K)

Log

H

R2 = 0.978

0.6

0.8

1.0

1.2

1.4

1.6

1.8

2.0

2.2

3.2 3.25 3.3 3.35 3.4 3.45 3.5 3.55 3.6 3.65

1000/T (K)

Lo

g H

a) c)

b) d)

R 2 = 0 .9 6 4

0.7

0.9

1 .1

1 .3

1 .5

1 .7

1 .9

2 .1

3.2 3.25 3.3 3 .35 3.4 3.45 3.5 3.55 3 .6 3.65

1000/T (K )

Lo

g H

R2 = 0.915

0.7

0.9

1.1

1.3

1.5

1.7

1.9

2.1

3.2 3.25 3.3 3.35 3.4 3.45 3.5 3.55 3.6 3.65

1000/T (K)

Lo

g H

R2 = 0.952

0.7

0.9

1.1

1.3

1.5

1.7

1.9

3.2 3.25 3.3 3.35 3.4 3.45 3.5 3.55 3.6 3.65

1000/T (K)

Log

H

R2 = 0.978

0.6

0.8

1.0

1.2

1.4

1.6

1.8

2.0

2.2

3.2 3.25 3.3 3.35 3.4 3.45 3.5 3.55 3.6 3.65

1000/T (K)

Lo

g H

a) c)

b) d)

67

Slope Intercept

Compound m S.D. b S.D. r 2

HCB a -3013 174 11.6 0.59 0.95

TC a -1524 158 6.58 0.54 0.95

CC a -1786 209 7.42 0.71 0.99

TN a -2068 284 8.44 0.97 0.91

p,p'-DDE -2043 240 8.37 0.82 0.96

a) HCB: hexchlorobenzene; TC: trans-chlordane; CC: cis-chlordane and TN: trans-nonachlor.

39.1 4.6 33

34.2 4 27

39.6 5.4 32

57.7 3.3 35

29.2 3 29

(kJ mol -1) S.D. (Pa m3 mol-1)

∆∆∆∆Hwa HLC at 25oC

Table 1. Parameters of log H = m/T + b and enthalpies of water-air transfer (∆∆∆∆Hwa).

3.2. Enthalpies of water-to-air exchange

Values of ∆Hwa (kJ mol-1) for the five chemicals are found in Table 1. The ∆Hwa for HCB (57.7 kJ mol-

1) compares favourably with 48 kJ mol-1 reported by ten Hulscher et al. (1992). Except for HCB, the ∆Hwa

values of the other pesticides (29.2 - 39.6 kJ mol-1) are lower than those for toxaphene (61.4 kJ mol-1) (Jantunen

and Bidleman, 2000) and hexachlorocyclohexanes (HCHs) (57.5 - 65.1 kJ mol-1) (Jantunen and Bidleman, 2000;

Sahsuvar et al., 2003), but greater than ∆Hwa for endosulfan-I (16.8 kJ mol-1) ( Rice et al., 1997), which is a

cyclodiene pesticide like the chlordane compounds. Staudinger and Roberts (2001) reviewed and compiled

experimental data for HLCs and ∆Hwa of compound classes including light hydrocarbons and chlorinated

hydrocarbons, polychlorinated biphenyls, pesticides and oxygenated chemicals (phenols, alcohols, acids, esters,

aldehydes, ketones) and nitrogen-containing compounds (alkyl nitrates and nitrogen heterocycles). The range of

∆Hwa across all compound classes was 8 - 93 kJ mol-1 with a mean of 47 kJ mol-1, while the range and mean for

eight pesticides (including endosulfan I and the HCHs, mentioned above) were 17 - 88 kJ mol-1 and 43 kJ mol-1.

Enthalpies of vapourization (∆Hvap) (kJ mol-1) for the compounds studied here are HCB = 68.5, TC = 80.7, CC =

82.0, TN = 85.5, p,p'-DDE = 87.2 (Hinckley et al., 1990). For all compounds, ∆Hwa is smaller than ∆Hvap, which

is expected since ∆Hwa = ∆Hvap - ∆Hsol and the enthalpy of solution in water (∆Hsol) is often positive for

hydrophobic organic chemicals (Paasivirta et al., 1999). Using the above equation, the ∆Hsol was calculated for

HCB. ∆Hsol is small (10.8 kJ mol-1) and similar to ∆Hsol for α- and γ-HCH (9.2 and 9.1 kJ mol-1) (Sahsuvar et al.,

2003) whereas ∆Hsol for the other compounds are larger and range from 45.9 - 51.5 kJ mol-1.

4. Application to air-water gas exchange in Lake Ontario

The net air-water gas exchange direction was estimated for TC, CC, TN and p,p’-DDE in Lake Ontario,

which lies on the Canada - U.S. border (44N, 76W). Paired air and surface water samples were collected from

the lake in July 1998 and analysed by capillary gas chromatography - electron capture negative ion mass

spectrometry using methods described elsewhere (Jantunen and Bidleman, 2003; Helm et al., 2003).

Concentration data from that study were used for the exchange estimates, except for the air concentration of p,p'-

DDE, which was taken from Buehler et al. (2000) for the same year. No parallel air and water data were

available for HCB so no air-water exchange determinations were made.

68

Method and Temperature, oC HCB TC CC TN Reference

bubble stripping

25 35 29 27 32 this study23 130 135 89 Atlas et al. 198225 71 Atlas et al. 1983

25 170 4.9a 4.9a Warner et al. 198720 41 ten Hulscher et al. 199220 49 Oliver et al. 1985

wetted wall column

23 8.8 Fendinger et al. 198925 24 Altschuh et al. 1999

fog chamber

23 5.9 Fendinger et al. 1989

calculated

vp/solubility, 25 16.5 11.3 49.2 Iwata et al. 1993Final Adjusted Values 65 6.8 5.7 Shen and Wania 2005.

a Reported as "chlordane", isomer not specified

4.2

4.2

33122

p,p'-DDE

Table 2. Comparison of Henry's law constants, Pa m3 mol-1

Water/air fugacity ratios were calculated from relationships given in Jantunen and Bidleman (2003), using

mean concentrations of the dissolved and vapour-phase compounds in water and air, air temperature, HLCs at the

water temperature and the gas constant R. Values of the fugacity ratio are <1, =1 and >1 for net deposition,

equilibrium and net volatilization conditions, respectively. Using the HLCs determined in this study, the

resulting fugacity ratios were 3.1 to 4.6 for the chlordanes and 11 for p,p’-DDE, indicating that these compounds

were undergoing net volatilization from Lake Ontario during July, 1998 (Table 3).

Figure 3 compares the fugacity ratios of trans-chlordane using different sets of literature HLC values (Table

2). Adjustments for temperature were done assuming the parameters of eq 5 determined in this work (Table 1).

Resulting fugacity calculations predict equilibrium or volatilization, depending on the reported HLC values

(Figure 3). The HLC of Iwata et al. (1993) predicts the same net exchange direction as this study and the

fugacity ratios are relatively close in magnitude. The FAV HLC from Shen and Wania (2005) predicts the same

net exchange direction as this study, but the magnitude is smaller (fugacity ratio = 1.5). Net volatilization is also

predicted from the HLC of Atlas et al. (1982) and the fugacity ratio (18) is much larger. Finally, using the HLCs

reported by Fendinger et al., (1988; 1989) result in a close-to-equilibrium fugacity ratio (0.8-1.2).

Similarly, using the HLCs of p,p'-DDE determined by ourselves, Altschuh et al. (1999) and Atlas et al. (1982)

predict net volatilization of p,p'-DDE from Lake Ontario, though with greatly different fugacity ratios, 11, 1.4

and 45, respectively.

69

Table 3: Air (Ca, pg m-3) and water (Cw, pg L-1) concentrations from Lake Ontario, July 1998 used to calculate fugacity ratios.

Compound Water Air Cw S.D. Temp (K) Ca S.D. Temp (K) log H HLC fw fa

TC 1.9 0.35 293 5 2.7 294 1.38 24 1.60E-07 4.20E-08CC 3.3 0.52 293 6.3 3.1 294 1.33 21 2.40E-07 5.30E-08TN 1.9 0.45 293 6.2 3.1 294 1.39 25 1.60E-07 5.20E-08

DDE 22 3 293 21a 16 294 1.41 25 1.90E-06 1.80E-07

a) Buehler et al., 2000.

11

3.84.63.1

fw/fa

Figure 3: Fugacity ratios for trans-chlordane calculated with the HLCs determined in this study in comparison to literature values. The error bars is derived from propagation of errors (as in Sahsuvar et al., 2003) for this study only.

5. Conclusions

HLCs for chlordane compounds, HCB and p,p'-DDE were determined as a function of temperature. These

are the first reported temperature-dependent values for chlordanes and p,p'-DDE and only the second ones for

HCB. Our HLCs for these compounds at 20-25oC are intermediate among literature values which vary by 1-2

orders of magnitude. As a consequence, air-water gas exchange estimates for chlordanes and p,p'-DDE in Lake

Ontario indicate either net volatilization, net deposition, or equilibrium conditions, depending on which set of

HLCs is used.

Although all researchers working in this field would like to believe that their HLCs are the "best", in fact it is

impossible to give an accurate description of gas exchange based on the present level of uncertainty. One way to

resolve the large discrepancy in reported HLCs is to carry out an adjustment technique to derive a set of

physicochemical properties which is thermodynamically consistent (Beyer et al., 2002). Properties included in

trans-Chlordane

0

1

2

3

4

5

6

7

8

This Study Iwata et al.1993.

Fendinger etal. 1989

Fendinger etal. 1988.

Mackay et al.1997.

Atlas et al1982

Shoeib andHarner, 2002.

fw/f

a

0.0

1

70

this exercise are the fundamental ones of solubility in water, octanol and air (vapour pressure) and the

partitioning properties HLC, Kow and Koa. This technique has been used in the evaluation and selection of the

optimum values for physicochemical properties for polychlorinated biphenyls (Li et al., 2003),

hexachlorocyclohexanes (Xiao et al., 2004) and organochlorine pesticides (Shen and Wania, 2005). The HLCs

determined here were about 2 times lower than the Shen and Wania FAV for HCB and 4-8 times higher than the

FAVs for the chlordane isomers and p,p'-DDE. HLC values for the three major HCH isomers, determined using

the same bubble stripping apparatus (Sahsuvar et al., 2003), agree within ~30% of FAVs (Xiao et al., 2004).

Recently a discussion, has taken place concerning the accuracy of the bubble stripping technique for

determining HLCs of highly hydrophobic chemicals. The discussion involved a potential artifact caused by

adsorption of chemical to the bubble surface, thereby enhancing its stripping rate from water. This has been

proposed by Goss et al. (2004) and disputed by Baker et al. (2004). The compounds included in this study have

the log Kow values HCB = 5.6, trans-chlordane = 6.3, cis-chlordane = 6.2 and p,p'-DDE = 6.9 (FAVs, Shen and

Wania, 2005). The Kow of p,p'-DDE is an order of magnitude greater than for HCB, but there is no great

difference in the hydrophobicities of HCB and the two chlordanes. Thus, it is difficult to understand why our

HLC value for HCB is lower than the FAV of Shen and Wania, whereas our HLCs for the other three compounds

are all higher. Further work is required to determine if differences in methodology are responsible for observed

variations in reported HLCs.

References Alaee, M., Whittal, R.M., Strachan, W.M.J. 1997. The effect of water temperature and composition on Henry’s law constant for various PAHs. Chemosphere 32, 1153-1164. Altschuh, J., Bruggermann, R., Santl, H., Eichinger, G., Piringer, O.G. 1999. Henry’s Law constants for a diverse set of organic chemicals: experimental determination and comparison of estimation methods. Chemosphere 39, 1871-1877. Atlas, E., Foster, R., Giam, C.S. 1982. Air-sea exchange of high molecular weight organic pollutants: laboratory studies. Environ. Sci. Technol. 16, 283-286. Atlas, E., Velasco, A., Sullivan, K., Giam, C.S. 1983. A radiotracer study of air-water exchange of synthetic organic compounds. Chemosphere 12, 1251-1258. Bamford, H., Poster, D.L., Baker, J.E. 1999. Temperature dependence of Henry’s law constants of thirteen polycyclic aromatic hydrocarbons between 4 oC and 13 oC. Environ. Toxicol. Chem. 18, 1905-1912. Bamford, H., Poster, D.L., Baker, J.E. 2000. Henry’s law constants of polychlorinated biphenyl congeners and their variation with temperature. J. Chem. Eng. Data 45, 1069-1074. Baker, J.E., Totten, L.A., Gigliotti, C.L., Offenberg, J.H., Eisenreich, S.J., Bamford, H.A., Huie, R.E., Poster, D.L. 2004. Response to comment on "Re-evaluation of air-water exchange fluxes of PCBs in Green Bay and southern Lake Michigan". Environ. Sci. Technol. 38, 1629-1632. Beyer, A., Wania, F., Gouin, T., Mackay, D., Matthies, M. 2002. Selecting internally consistent physicochemical properties of organic compounds. Environ. Toxicol. Chem. 21, 941-953.

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Breitner, W.A., Baker, J.M., Kostinen, W.C. 1998. Direct measurement of Henry’s Law constants for S-ethyl N,N-di-n-propylthiocarbamate. J. Agric. Food Chem. 46, 1624-1629. Buehler, S., Hafner, W., Basu, I., Audette, C.V., Brice, K.A., Chan, C.H., Froude, F., Galarneau, E., Hulting, M.L., Jantunen, L., Neilson, M., Puckett, K., Hites, R.A. 2000. Atmospheric deposition of toxic substances to the Great Lakes: IADN results through 1998, Environment Canada and the United States Environmental Protection Agency, ISBN 0-662-31219-8. Fendinger, N.J., Glotfelty, D.E. 1988. A laboratory method for the experimental determination of air-water Henry’s law constants for several pesticides. Environ. Sci Technol. 22, 1289-1293. Fendinger, N.J., Glotfelty, D.E. 1989. Comparison of two experimental techniques for determining air/water Henry’s law constants. Environ. Sci. Technol. 23, 1528-1531. Gautier, C., Le Calvé, S., Mirabel, P. 2003. Henry's law constants measurements of alachlor and dichlorvos between 283 and 298 K. Atmos. Environ. 37, 2347-2353. Goss, K-U., Wania, F., McLachlan, M.S., Mackay, D., Schwarzenbach, R.P. 2004. Comment on "Re-evaluation of air-water exchange fluxes of PCBs in Green Bay and southern Lake Michigan". Environ. Sci. Technol. 38, 1629-1632. Gossett, J.M. 1987. Measurement of Henry’s Law constants for C1 and C2 chlorinated hydrocarbons. Environ. Sci. Technol. 21, 202-208. Hansen, K. C., Zhou, Z., Yaws, C. L., Aminabhavi, T. M. 1993. Determination of Henry’s law constants of organics in dilute aqueous solutions. J. Chem. Eng. Data 38, 546-550. Hinckley, D.A., Bidleman, T.F., Foreman, W.T., Tuschall, J.R. 1990. Determination of vapor pressures for nonpolar and semipolar organocompounds from gas chromatographic retention data. J. of Chem. Eng. Data 35: 232-237. ten Hulscher, Th. E.M., van der Velde, L.E., Bruggeman, W.A. 1992. Temperature dependence of Henry’s law constants for selected chlorobenzenes, polychlorinated biphenyls and polycyclic aromatic hydrocarbons. Environ. Toxicol. Chem. 11, 1595-1603. Iwata, H., Tanabe, S., Sakai, N., Tatsukawa, R. 1993. Distribution of persistent organochlorine pollutants in oceanic air and surface seawater and the role of ocean on their global transport and fate. Environ. Sci. Technol. 27, 1080-1098. Jantunen, L.M., Bidleman, T.F. 2000. Temperature dependent Henry’s law constant for technical toxaphene. Chemosphere, Global Change Science 2, 225-231. Jantunen, L.M.; Bidleman, T.F., 2003. Air-water gas exchange of toxaphene in Lake Superior. Environ. Toxicol. Chem. 22, 1229-1237. Kucklick, J.R., Hinckley, D.A., Bidleman, T.F. 1991. Determination of Henry’s law constants for hexachlorocyclohexanes in distilled water and artificial seawater as a function of temperature. Mar. Chem. 34, 197-209. Li, N., Wania, F., Lei, Y.D., Daly, G.L. 2003. Comprehensive and critical compilation, evaluation and selection of physical chemical property data for selected polychlorinated biphenyls. J. Phys. Chem. Ref. Data 32, 1535-1590. Mackay, D., Shiu, W.Y., Sutherland, R.P. 1979. Determination of air-water Henry’s law constants for hydrophobic pollutants. Environ. Sci. Technol. 13, 333-337.

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Murphy, T.J., Mullin, M.D., Meyer, J.A. 1987. Equilibration of polychlorinated biphenyls and toxaphene with water and air. Environ. Sci. Technol. 21, 155-162. Oliver, B.G. 1985. Desorption of chlorinated hydrocarbons from spiked and anthropogenically contaminated sediments. Chemosphere 14, 1087-1106. Paasivarta, J., Sinkkonen, S., Mikkelson, P., Rantio, T., Wania, F. 1999. Estimation of vapour pressures, solubilities and henry’s law constants of selected persistent organic pollutants as functions of temperature. Chemosphere 39, 811-832. Rice, C.P., Chernyak, S.M., McConnell, L.L. 1997. Henry’s law constant for pesticides measured as a function of temperature and salinity. J. Agric. Food Chem. 45, 2291-2298. Sahsuvar, L., Helm, P.A., Jantunen, L.M., Bidleman, T.F. 2003. Henry’s Law constants for α-, β- and γ-hexachlorocyclohexanes (HCHs) as a function of temperature and gas exchange in arctic regions. Atmosph. Environ. 37, 983-992. Shen, L., Wania, F. 2005. Compilation, evaluation and selection of physical-chemical property data for organochlorine pesticides. J. Chem. Eng. Data 50, 742-768. Shepson, P.E., Mackay, E., Muthuramu, K. 1996. Henry’s law constants and removal processes for several atmospheric α-hydroxy alkyl nitrates. Environ. Sci. Technol. 30, 3618-3623. Staudinger, J., Roberts, P.V. 2001. A critical review of Henry’s law constant temperature dependence relations for organic compounds in dilute aqueous solutions. Chemosphere 44, 561-576. US EPA, Part 5 Chemical-Specific Parameters from Superfund Chemical Matrix, 9345.1-21 EPA 540/R 96/028, June 1996, http://www.epa.gov/superfund/resources/soil/part_5.pdf. Warner, H., Cohen, J.M., Ireland, J.C. 1987. Determination of Henry’s Law Constants of Selected Priority Pollutants. Environmental Protection Agency, Cincinnati, OH, Water Engineering Research Lab, EPA/600/D-87/229. Xiao, H., Li, N., Wania, F. 2004. Compilation, evaluation and selection of physical-chemical property data for α-, β- and γ-hexachlorocyclohexane. J. Chem. Eng. Data 49, 173-185. Yin, C., Hassett, J.P. 1986. Gas partitioning approach for laboratory and field studies of mirex fugacity in water. Environ. Sci. Technol. 20, 1213-1217.

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Paper 2

Toxaphene, Chlordane, and Other Organochlorine Pesticides in Alabama Air

Liisa M.M. Jantunen 1, Terry F. Bidleman 1, Tom Harner1, William J. Parkhurst2

Environmental Science and Technology, 2000, 34, 5097-5105.

1 Environment Canada, 4905 Dufferin Street, Downsview, Ontario, M3H 5T4, Canada 2 Tennessee Valley Authority, Muscle Shoals, Alabama, USA. Contributions: Tom Harner and Terry Bidleman made arrangements for air sampling in Alabama, and the samples were taken by William Parkhurst at Tennessee Valley Authority. Liisa Jantunen prepared, extracted, cleaned up and analysed the samples. Liisa did the data interpretation and wrote the paper. Terry Bidleman secured funding and provided scientific guidance during every step during this project.

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Abstract

Air samples were collected in Alabama in January-October 1996 and again in May 1997, to determine the

seasonal variations of toxaphene and other OC pesticides (chlordanes, DDTs, dieldrin and HCHs). Log partial pressure

versus 1/T plots for γ-HCH, dieldrin, chlordanes and their metabolites showed significant relationships, where those for

toxaphene, heptachlor and p,p’-DDE did not. The chromatographic profile of toxaphene congeners in air and soil

showed depletion of certain labile congeners, notably B8-531 and B8-806/809; other persistent congeners (B8-1412 and

B9-1679) were enriched in air samples. Concentrations of toxaphene in Alabama air were 6-40 time higher than in the

Great Lakes region and chlordane in Alabama air exceeded Great Lakes concentrations by 3-9-fold. This suggests that

transport from the southern U.S. is a continuing source of toxaphene and chlordane to the Great Lakes. Levels of HCHs

and dieldrin in Alabama were similar to those in the Great Lakes region. Patterns of 8-chlorinated bornanes in air more

closely resemble residues in soil than the technical toxaphene standard. Enantiomer ratios of chlordane in air were

nearly racemic, indicating that their source is probably evaporation from termiticide usage rather than soils, which

contain non-racemic chlordanes.

Introduction

Organochlorine pesticides (OCs) were at one time widely used in North America. Compounds included

toxaphene, chlordane, aldrin, dieldrin, heptachlor and DDT. Although these chemicals were banned 10-28 years ago

they are still present in ambient air in the Great Lakes (Cortes et al., 1998) and the Arctic (Halsall et al., 1998). Air

monitoring in the early to mid 1990s by the Integrated Atmospheric Deposition Network (IADN) found a decrease in the

concentrations of OCs around the Great Lakes and these time trends were used to project virtual elimination dates (when

concentrations fall below detectability by today’s methods) of 2010-2070 (Cortes et al., 1998). Since these pesticides

have not been used for decades, some mechanisms are maintaining current ambient levels. These may include transport

from countries south of the U.S.A. border where some OCs are currently applied (Alegria et al., 2000) and re-emissions

from agricultural soils in past usage regions. Spencer et al. (1996) measured DDT in air above the soils of a farm in

California where DDT was last used 23 years previously and found 9.2-16.9 ng/m3 total DDTs in February and

September. These levels are 2-3 orders of magnitude higher than those at IADN stations (Cortes et al., 1998). Elevated

concentrations of OCs were also found in air above agricultural soils in British Columbia (Finizio et al., 1998).

High episodes of OCs in air sampled in southern Ontario were associated with air masses that arrived from the

southern U.S, versus lower concentrations during northerly transport (Hoff et al., 1992a, b). Ambient air samples taken

in South Carolina in 1994-95, showed levels of chlordanes and toxaphene that were an order of magnitude higher than

those found in the Great Lakes region (Bidleman et al., 1998). No relationship of toxaphene concentrations in South

Carolina to air transport direction was found, suggesting that the toxaphene was volatilized from regional soils.

The purpose of the study was to measure airborne OC pesticides in Alabama to determine whether the southern

U.S. is a potential source region for transport to the Great Lakes. Alabama was selected because toxaphene usage there

was the highest in southern states in 1980 (Voldner and Schroeder, 1989). Residues of toxaphene and other OCs were

recently reported for Alabama agricultural soils (Harner et al., 1999). In a companion paper (Harner et al., 2000),

75

measured atmospheric concentrations are compared to those predicted from fugacity based volatilization model, which

uses soil residues as drivers.

Experimental Methods

Sample Collection and Preparation

Air samples were collected in northern Alabama at Tennessee Valley Authority (TVA) headquarters. The site

(34o48’N, 87o 40’W) is a 435 000 acre reservation abutting the towns of Muscle Shoals and Florence, and is surrounded

by farmland. Samples were collected from January to October 1996 and in May 1997. In most months, 1-2 samples

were collected except for August 1996 and May 1997 when 6 and 7 samples per month were taken (Table 1). A PS-1

(Wedding) sampler was used, consisting of a 10.2 cm diameter glass fiber filter (Whatman, 0.1 µm nominal cutoff)

followed by two polyurethane foam (PUF) traps (6.8 cm diameter x 4.2 cm) in a glass cylinder. Indoor air was also

sampled for chlordanes in five private homes in Muscle Shoals, using a low volume (30 L/min) collector and smaller

PUF traps (5cm diameter x 3cm thick). PUF traps and filters were cleaned before use as previously described by

Billings and Bidleman, 1980. Samples were collected over a ~24 h period with air volumes ranging from 334-465 m3.

Filters were refluxed in dichloromethane and PUFs were Soxhlet extracted in petroleum ether overnight. The extracts

were concentrated by rotary evaporation, blown down with a gentle stream of nitrogen and exchanged into isooctane.

Extracts were cleaned up and fractionated on an alumina-silicic acid column to separate PCBs from the chlorinated

pesticides. The column consisted of 3 g silicic acid (deactivated with 3% water), overlaid with 1 g neutral alumina

(deactivated with 6% water) and topped with 1 cm anhydrous granular Na2SO4 (Keller and Bidleman, 1984). The

column was prewashed with 30 mL dichloromethane then 30 mL petroleum ether. The sample was eluted with 20 mL

petroleum ether (fraction 1, containing the PCBs, heptachlor, a portion of the DDE and o,p’-DDT and a small amount of

the Cl-8 toxaphenes) and then 30 mL dichloromethane (fraction 2, contains the remainder of the chlorinated pesticides).

Fractions were transferred to iso-octane and adjusted to ~1.0 mL for analysis. If necessary, additional cleanup was done

by shaking the extract with 18M H2SO4 (omitted for analysis of the oxygen-containing OCs, these included dieldrin,

heptachlor exo-epoxide and oxychlordane).

Analysis

Pesticides were determined by gas chromatography-negative ion mass spectrometry (GC-NIMS) on a Hewlett

Packard 5890 GC-5989B MS Engine with methane at a nominal pressure of 1.0 Torr. The column used for quantitative

analysis was a DB-5 (J&W, 30 m x 0.25 mm i.d., 0.25 µm film thickness) operated at a helium carrier gas flow of 40

cm/s. The temperature program was: initial temperature 90oC, 15o/min to 160o, 3.5o/min to 210o, hold 1.0 min, 20o/min

to 260o hold 5.0 min. Sample volumes of 2 µL were injected splitless (split opened after 1.5 min). Other temperatures

were: injector and transfer line 250o, ion source 150o and quadrupole 100o. Mirex was added to extracts prior to

injection as the internal standard. Random samples were checked for native mirex and found negative. Toxaphene was

quantified as the sum of the 7-Cl, 8-Cl and 9-Cl homologues by monitoring ions 343/345, 379/381 and 413/415

(Jantunen and Bidleman, 1998). Other compounds sought (abbreviations and ions monitored; target/qualifying) were:

hexachlorocyclohexanes (HCHs; 255/257), trans- and cis-chlordane (TC and CC; 412/410), trans- and cis-nonachlor

76

(TN and CN; 444/446), heptachlor (HEPT; 300/302), oxychlordane (OXY; 422/420), heptachlor exo-epoxide

(HEPX; 388/386), DDT (246/248/250), DDE (316/318) and mirex (404). Quantification of all OCs except total

toxaphene was carried out against five standards that spanned the concentration range of the samples, using the HP MS

Chemstation software.

Two methods were used to quantify total toxaphene, based on single and multiple response factors (SRF and

MRF) (Jantunen and Bidleman, 1998). The SRF method uses the sum of areas of the Cl-7, Cl-8 and Cl-9 homolog

groups of the standard and sample. In the MRF approach, peaks in the technical mixture are assigned mass percentages

of the total by using GC-flame ionization detection. NIMS response factors are then determined from peak areas, their

mass percentages and the total quantity of toxaphene standard injected. The NIMS response factors are used to quantify

individual peaks in the sample and results are summed to yield a value for total toxaphene. A similar method has also

been described by Shoeib et al., (1999). Single chlorobornane congeners were quantified versus pure standards as for

the other OC pesticides.

Two different chiral columns were used for the enantiomeric analysis: Beta-DEX 120 (20% permethylated β-

cyclodextrin in polydimethylsiloxane, 30 m x 0.25 mm i.d., 0.25 µm film thickness, Supelco) and BGB-172 (20% tert-

butyldimethylsilylated β-cyclodextrin in OV-1701, 30 m x 0.25 mm i.d., 0.25 µm film thickness, BGB Analytik AG,

Switzerland). The Beta-DEX temperature program was: initial temperature 90oC, hold for 1.0 min, 15o/min to 130o,

1.0o/min to 210o, 20o/min to 230o hold for 5.0 min. The BGB-172 program was: 90o, hold for 1.0 min, 15o/min to 140o,

1.0o/min to 190o, hold for 2.0 min, 20o/min to 240o hold for 5.0 min. Both columns were operated at a He carrier gas

flow of 40 cm/s. Other conditions were: splitless injection (split opened after 1.5 min), injector 220o, transfer line 220o,

ion source 150o and quadrupole 100o. Enantiomers of α−HCH, HEPT and HEPX were separated on the BGB-172

column. The Beta-DEX column resolved the enantiomers of TC and CC and was used as a confirmation column for

α−HCH since the elution order of the enantiomers is reversed from the order on BGB-172 (Jantunen and Bidleman

1998).

Analytical standards for the cyclodienes and HCHs were purchased from Supelco (Bellefonte, PA) in a 20

component mixture; added to this mixture were technical toxaphene and individual standards of trans- and cis-nonachlor

and oxychlordane (Supelco). Single chlorobornane congeners were purchased from Axact Standards (Commack, NY).

The DDTs were purchased individually from Supelco and combined in a separate standard. Technical chlordane and

toxaphene standards were obtained from the U.S. Environmental Protection Repository for Pesticides and Industrial

Chemicals (EPA, Research Triangle Park, NC, 1988 and 1990). An additional technical chlordane standard was

obtained from Radian Corporation (1985). All solvents were chromatographic quality.

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Table 1: OCs in Alabama Air, January to October 199 6 and May 1997 (pg/m 3), + indicates standard deviation.

Month n αααα-HCH γγγγ-HCH TC CC TN CN TC:CC:TN HEPT HEPX OXY p,p'-DDE Dieldrin Toxaphene

January 1 29 6 7 3 3 0.1 1.00:0.40:0.36 4 4 1 1 6 8

February 2 180-341 27-34 23-24 9-11 9-10 0.8-0.9 1.00:0.43:0.40 5-9 5-9 1-3 4-7 10-15 172-224

March 1 124 40 26 11 10 1 1.00:0.42:0.39 7 7 1 4 12 135

April 1 113 55 14 6 7 0.5 1.00:0.45:0.51 8 8 3 2 17 58

May 1 7 39+20 52+29 27+15 17+11 17+11 2.3+2.7 1.00:0.63:0.66 21+10 22+22 8.5+8.6 6+13 64+66 113+64

June 2 72-98 77-110 73-95 45-55 41-49 9-10 1.00:0.60:0.54 43-49 29-43 15 53-92 55-89 474-611

July 2 22-65 61-72 72-105 42-53 41-44 5-6 1.00:0.55:0.50 35-46 18-24 8-10 2 50-74 121-545

August 7 48+20 57+28 82+46 47+25 38+18 4+2 1.00:0.58:0.49 54+44 23+10 10+4 1 58+31 194+105

September 1 86 77 83 40 36 5 1.00:0.48:0.43 45 21 8 7 46 284

October 1 89 23 35 18 18 2 1.00:0.51:0.50 na 14 7 1 30 75

Annual Mean 92 50 47 25 23 3.1 1.00:0.56:0.52 26 16 6.5 10 38 176Standard Deviation 68 26 33 19 16 2.9 0:0.09:0.13 20 10 4.6 22 25 151

HCH: hexachlorocyclohexane, TC: trans-chlordane, CC cis-chlordane, TN: trans-chlordane, HEPT: heptachlor, HEPX: heptachlor epoxide and OXY: oxychlordane.1: May 1966 and 1997 were averaged

78

Figure 1: Concentrations in pg/m3 a) γ-HCH, b) heptachlor (HEPT) and cis-nonachlor (CN), c) trans-chlordane (TC) and trans-nonachlor (TN), d) cis-chlordane (CC) and heptachlor exo-epoxide (HEPX), e) dieldrin and oxychlordane (OXY), f) p,p’-DDE, g) total toxaphene. Scales on the right pertain to the open bars. Results and Discussion Quality Control

Jan Feb Mar Apr May Jun Jul Aug Sep Oct0

50

100

150

200

a-HCH g-HCH


20

40

60

80

0

2

4

6

8

10

12

HEPT CN


20

40

60

80

100 TC TN


10

20

30

40

50

60 CC HEPX


20

40

60

80

0

5

10

15

20

Dieldrin OXY


100

200

300

400

500

600

SRFMRF


20

40

60

80 p,p'-DDEpg/m

3

Toxaphene

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Recovery experiments were done by spiking a clean PUF with 166-253 ng single-component pesticides and

254 ng toxaphene. Average recoveries (n=3) were: toxaphene 96±3%, chlordanes 92±1%, DDTs 100±5%, HCHs

79±6% and dieldrin 94±4%. Clean PUFs (n=4) were extracted and processed in the same manner as the samples to

provide blank values. No peaks were seen above the noise level. The lowest concentration of standard injected was 0.8

pg/µL for single-component pesticides and 12 pg/µL for toxaphene. The resulting peaks had average signal to noise

ratios of 15-50, assuming that one-tenth of these concentrations would be detectable for most compounds, the detection

limits were approximately < 0.2 pg/m3 air for single-component pesticides and <3.5 pg/m3 for toxaphene, based on a

350-m3 air volume.

The front and back PUFs were analyzed separately to assess the breakthrough by comparing the amount found

on each. During the warm months breakthrough was significant for α-HCH, when α-HCH on the back PUF amounted

to 50% or more of the front PUF value in most samples. Thus, results here for α-HCH are reported as the sum of both

PUFs and are likely to be lower limits. Breakthrough was less for other OCs, (expressed as back/front amounts) e.g. 1-

20% for γ-HCH, 0.1-1.5% for TC. In the cooler months breakthrough was not found except for α-HCH.

Concentrations of OCs in Alabama Air. PUFs plugs and filters were analyzed separately, and negligible quantities of

the OCS were retained by the filter. Percentages found on the filters were ≤2% for the chlordanes, ≤5% for DDE and

<1% for the other OCs.

Toxaphene: Toxaphene is a complex mixture of several hundred compounds (Hainzl et al., 1994), which was mainly

used on cotton and soybeans in the southern U.S. and also to a lesser extent in other states. The peak usage of toxaphene

was 25 x 106 kg/yr in 1972 (Voldner and Schroeder, 1989). Most registrations of toxaphene were canceled in 1982, but

remaining stocks were applied through 1986.

The range and mean ± standard deviation concentrations (pg/m3) of total toxaphene in air were 8-611 and

176±151 for the SRF method and 6-530 and 142±137 for the MRF method (Table 1 and Figure 1). On average, the

MRF yielded results 20% lower than the SRF and toxaphene residues quantified by the two methods were correlated by

SRF = 0.947 MRF – 15.5. The two methods were also used to quantify toxaphene over Lake Superior and toxaphene in

arctic water; in both cases the results of the MRF and SRF calculations were not significantly different (Jantunen and

Bidleman 1998; 2003). Toxaphene residues quantified by the two methods were correlated by SRF = m⋅MRF + b,

r2=0.91, m =0.947 and b =-15.5. The annual mean concentration in Alabama air calculated by the SRF method is

similar to the mean for July 1994 - January 1995 in Columbia, South Carolina, 189±107 pg/m3 (Bidleman et al., 1998).

Qualifying peaks matching in retention time to two persistent congeners were quantified against pure standards.

The two congeners were 2-endo,3-exo,5-endo,6-exo,8,8,10,10-octachlorobornane and 2-endo,3-exo,5-endo,6-

exo,8,8,9,10,10-nonachlorobornane. The octachloro compound is also known as Parlar 26 (P26; Frenzen et al., 1994),

T2 (Stern et al., 1992) and B8-1413 (Andrews and Vetter, 1995) and the nonachloro compound as P50, T12, B9-1679

and Toxicant Ac (Saleh, 1991). Other Parlar congeners

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Figure 2: Chromatograms of total toxaphene and Cl-7 to Cl-9 homologue groups. Top = standard, middle = soil and bottom = air. Peak 1 = B8-1413, 2 = B8-1412, 3 = B8-531, 4 = B8- 1945, 5 = B8-806/809, 6 = B8-2229, 7 = B9-1679 and 8 = B9-2206.

were identified (but not individually quantified) by their position in the toxaphene pattern (Buser et al., 1998) (Figure 2).

A persistent octachlorobornane (B8-1412; Klobes et al., 1997) which elutes just after B8-1413 was identified by using a

standard supplied by Walter Vetter, University of Jena, Germany. Although peaks matching B8-1413 and B9-1679

were quantified as these compounds, they may not be single compounds in air samples. This was shown by Shoeib et al.

(2000), who used multidimensional GC-ECD to examine the composition of toxaphene peaks in air samples. In the

Alabama air samples, B8-1413 and B9-1679 concentrations averaged 3.2 pg/m3 and 3.6 pg/m3, respectively. We found

B8-1413 = 0.52% and B9-1679 = 1.2% of technical toxaphene standard, which compare well to the percentages

reported by Shoeib et al. (1999) of 0.49% and 1.51%. As percentages of total toxaphene in air B8-1413 and B9-1679

percentages in air ranged from 0.2-8.8% and 0.9-12% and averaged 1.9±2.0% and 3.3±3.4%. This shows significant

(p<0.01) enrichment of both congeners in air compared to the percentages in the technical standard. Vetter and Scherer

(1999) showed that B8-1413 and B9-1679 have very stable structures in comparison to other polychlorinated bornanes

due to the placement of the chlorines on the bornane structure. Even though they account for only minor proportions of

the technical toxaphene they are the dominant congeners in marine mammals and human milk (Stern et al., 1992).

In 1996, Harner et al., (1999) collected soil samples from cotton fields in Alabama for analysis of toxaphene

and other OCs. Figure 2 shows profiles of total toxaphene and the Cl-7 to Cl-9 homolog groups for soil, air and the

technical standard. The soil somewhat reflects the technical mixture by being dominated by the mid-range congeners

whereas, air samples are enriched in the lighter earlier-eluting congeners. Both the soil and air profiles show evidence of

transformations from the EPA technical toxaphene standard, especially for a group of peaks in the Cl-8 and Cl-9

homologs (Figure 2). Peaks 3-6 in the Cl-8 homolog (Figure 2) elute within a minute of each other on a DB-5 column,

Air

Soil

Std

Total Toxaphene

Cl-8 homolog

Cl-9 homolog

Cl-7 homolog

4 6

5

46

5

4

65

7

7

7

8

8

8

3

3

3

12

21

21

81

so their vapor pressures are similar and therefore differences in proportions of the peaks are probably not due to

volatility. The pattern of these three peaks is similar in the soil and air but different from the technical product that

would have been applied. This implies that the toxaphene in ambient air is probably the result of volatilization of aged

residues rather than long range transport of newly applied pesticide. Peak 3 and 5 are depleted in the soil and air

compared to the standard. The major components of these peaks are B8-531 and B8-806/809 (Shoeib et al., 2000)

which are rapidly degraded in sewage sludge (Buser et al., 1998). Shoeib et al. (1999, 2000) also found similar

depletions in air samples from Point Petre, on the north shore of Lake Ontario.

Chlordane-Related Compounds: Technical chlordane was used mostly as an agricultural pesticide on corn and citrus,

for home lawns and gardens and as a termiticide in house foundations. The most abundant components in the technical

mixture are trans-chlordane, cis-chlordane, trans-nonachlor, β-chlordene and heptachlor (Dearth and Hites, 1991; PHS,

1992a). Technical heptachlor was also used as a termiticide, alone or in combination with chlordane (PHS, 1992).

Chlordane and heptachlor registrations in the U.S. were cancelled for all uses except termite control in 1983 and de-

registered for all uses in 1988.

The chlordane-related compounds sought in Alabama air were TC, CC, TN, CN, HEPT and the metabolites

OXY and HEPX. Atmospheric levels averaged (pg/m3): TC 47±33; CC 25±19; TN 23±16; CN 3.1±2.9; HEPT 26±20;

HEPX 16±10; and OXY 6.5±4.6. Much higher concentrations of chlordanes were found in the indoor air of five homes

in the Muscle Shoals area, averaging (pg/m3) TC 30 000 ± 28 000, CC 18 000 ± 19 000 and TN 9 000 ± 10 000.

The average proportions of TC:CC:TN in air samples were as follows: ambient air = 1.00 : 0.56: 0.52, and

indoor air = 1.00 : 0.61 : 0.29 (Table 2), CN was not included because concentrations were not determined for the soil

nor the indoor air samples. The agricultural soils of northwestern Alabama, where the air samples were taken, were

relatively low in chlordane. Geometric mean concentrations of TC, CC, and TN in the region were 0.057, 0.047 and

0.067 ng/g dry weight, compared to 0.49, 0.54 and 0.86 for the state as a whole (Harner et al., 1999). The geometric

mean was chosen because the concentrations in soil appear to be log-normally distributed (Harner et al., 1999). Lower

soil concentrations in the air sampling regions have not been explained but may be due to a large number of soil sampled

in this region compared to the rest of the state. The proportions of the chlordanes were also different; TC:CC:TN = 1.00

: 0.82 : 1.18 in northwest Alabama and 1.00 : 1.10 : 1.76 in the entire state (calculated from the geometric mean

concentrations, see above). The TC:CC:TN proportions in the EPA and Radian technical chlordane mixtures,

determined by GC-NIMS analysis versus single-component standards, were virtually identical: EPA = 1.00 : 0.85 : 0.42

and Radian = 1.00 : 0.86 : 0.41.

82

Table 2: Ratios of chlordanes compounds in air, soil and technical chlordane.

TC:CC:TN

soil, NW ALa 1.00:0.82:1.18

soil, all ALa 1.00:1.10:1.76technical chlordane 1.00:0.85:0.42ambient air, NW AL 1.00:0.56±0.09:0.52±0.13indoor air, Muscle Shoals 1.00:0.61±0.07:0.29±0.09

predicted ambient airb

from average soil residues, NW AL 1.00:0.59:0.65from average soil residues, all AL 1.00:0.76:0.81from technical chlordane 1.00:0.61:0.23

a Reference 10. b Using relative vapor pressures at 25oC TC:CC:TN = 1.00:0.72:0.55 (30).

Potential sources of chlordanes in Alabama air include volatilization from local and regional soils, release from

homes treated for termite control, and long-range transport. The expected proportions of chlordanes in air arising from

volatilization from soil or technical chlordane are calculated from the proportions in these sources using the relative

vapor pressures of compounds at 25oC, TC:CC:TN = 1.00 : 0.72 : 0.55 (Hinckley et al., 1990). Results are summarized

in Table 2. The proportions in the indoor air samples correspond closely to those in technical chlordane after

accounting for differences in volatility. The CC:TC ratios found in ambient air (0.56), indoor air (0.61) and estimated

from volatilization of technical chlordane (0.60) or soil residues in northwest Alabama (0.59) were all quite similar, but

lower than the CC:TC ratio estimated from volatilization of the state-averaged soil residues (0.76). Ratios of TN:TC

predicted from volatilization of soil residues in either northwest Alabama (0.65) or the entire state (0.81) were greater

than those in the indoor air (0.29) or from volatilization of technical chlordane (0.23).

These results do not clearly differentiate soil versus termiticide sources of TC and CC. However, chiral

analysis of the chlordanes suggests that volatilization from soil is not the main source and that termiticide emissions may

be more important (see below). The relatively high concentration of HEPT relative to HEPX in ambient air (Table 1)

also suggests a termiticide source. Furthermore, a soil-air exchange model indicates that chlordane emissions from

Alabama agricultural soils cannot support observed ambient air concentrations (Harner et al., 2000).

HCHs: Technical HCH is a mixture of several isomers: α-HCH (60-70%), β-HCH (5-12%), γ-HCH (10-15%) and

several other minor isomers (Iwata et al., 1993). Technical HCH was used in the U.S.A. until 1978, when it was replace

by the purified active isomer γ-HCH, also known as lindane. Lindane is still registered for restricted applications in

Canada and the U.S.A. (Walker et al., 1999). The two isomers sought in each sample were α- and γ-HCH.

Concentrations in Alabama air (pg/m3) averaged: α-HCH >92 and γ-HCH 50±26. The α-HCH is given as a lower limit

because of breakthrough losses for most samples (see Quality Control section). γ-HCH concentrations show a peak in

June and lower concentrations in the spring and fall; this is probably due to currently usage and a temperature-dependant

desorption mechanism from the soil. The α/γ-HCH ratio is > 1.0 (even considering the lower limit for α-HCH) for all

83

samples, implying that there is a significant historical background of technical HCH with additions of lindane through

current usage.

DDTs: DDT was first used to control disease spreading insects and then as a multipurpose insecticide. The peak

production of DDT in the U.S. was 82 million kg in 1962 and it was de-registered in 1972 except for public health

emergencies (PHS, 1992b). Only the degradation product, p,p’-DDE was found in all samples (Table 1, Figure 1).

Concentrations ranged from 0.25 - 91 pg/m3, with an average of 10±22 pg/m3. A soil volatilization model estimated that

the observed atmospheric levels of p,p’-DDE could be sustained by soil emissions (Harner et al., 2000). The o,p’-DDT

was detected in 11/25 (0.2-10.46 pg/m3), p,p’-DDT in 3/25 (0.36-1.36 pg/m3), o,p’-DDD 15/25 (0.1-4.6 pg/m3) and

o,p’-DDE in 14/25 (0.1-0.28 pg/m3) samples. The geometric mean concentrations of p,p’-DDT and p,p’-DDE in

agricultural soils of northwest Alabama were 12.4 and 17.5 ng/g dry weight; p,p’-DDT:p,p’-DDE ratio calculated from

the geometric mean concentrations = 0.71 (Harner et al., 1999). As for the chlordanes, geometric mean concentrations

were reported since DDT residues in soil varied by orders of magnitude (Harner et al., 1999).

Assuming that the p,p’-DDE in air comes from soil emissions, the expected p,p’-DDT concentration in air from

soil emissions would be only ~1 pg/m3, since the vapor pressure of p,p’-DDT at 25o is 15% of that for p,p’-DDE

(Hinckley et al., 1990). A value of 1 pg/m3 is close to the measurement value (when p,p’-DDT was detected) of 0.36-

1.4 pg/m3.

Dieldrin: Dieldrin enters the environment through direct application or the use of aldrin that is quickly transformed into

dieldrin in the environment. Both aldrin and dieldrin were used heavily in the 1950s to 1970s to combat insects on corn,

cotton and citrus crops, and also used as termiticides. The peak usage of aldrin was in 1966 and both dieldrin and aldrin

were de-registered in 1987 (PHS, 1992c). Aldrin was not sought in this study but dieldrin was found in all samples, at

concentrations ranging from 6-170 pg/m3, with a mean of 38±25 pg/m3. Atmospheric levels of dieldrin are still fairly

high considering the pesticide has not been used in agriculture since the late 1970s. Dieldrin concentrations in air at an

inland station in Belize, Central America were an order of magnitude higher than those found in Alabama and aldrin was

also high in the Belize samples (Alegria et al., 2000). This suggests that out-of-the-country sources may contribute to

atmospheric levels of dieldrin in Alabama.

Comparison to Great Lakes Regions

Average concentrations of OCs in ambient air from the southern U.S. and the Great Lakes regions are compared in

Table 3. The Great Lakes data are from two IADN stations (Eagle Harbor, Lake Superior and Sturgeon Point, Lake

Erie) (Cortes et al., 1998 ), other papers (Shoeib et al., 1999; Glassmeyer et al., 1999) and measurements made by

ourselves on cruises of Lake Superior in August 1996 and May 1997 (Jantunen et al., 2000). With an annual mean of

176 pg/m3, toxaphene in Alabama air was 10-20 times higher than in air over the Great Lakes. Toxaphene at Eagle

Harbor on Lake Superior in 1996-97 averaged 6.6 pg/m3 including one high value of 63 pg/m3 (Glassmeyer et al.,

1999). Mean concentrations measured from shipboard over Lake Superior were 28 pg/m3 in August and 11 pg/m3 in

May (Jantunen et al., 2000). Toxaphene averaged 3.8 pg/m3 at Point Petre on Lake Ontario in 1995-1997 (Shoeib et al.,

1999). Results from this study and the elevated levels found in South Carolina in 1994-1995 (Bidleman et al., 1998)

84

(Table 2) suggest that the transport from the southern U.S. may be a continuing source of toxaphene to the Great Lakes.

Hoff et al., (1992) used back trajectories from southern Ontario to investigate the origin of air masses. They concluded

that air masses with elevated OC pesticides were associated with transport from the lower United States or further south.

Voldner and Schroeder (1989) modeled toxaphene transport based on estimated usage patterns in the southern U.S.

They showed that toxaphene emitted from southern U.S. was transported and deposited in the Great Lakes at levels

similar to observed deposition patterns. Toxaphene was also transported along the northern U.S. coast and out over the

North Atlantic.

Chlordanes were also higher in the southern U.S. air than in the Great Lakes region. The Σ−chlordanes

(TC+CC+TN) averaged 95 pg/m3 in Alabama and 180 pg/m3 in urban South Carolina (Bidleman et al., 1998), compared

to 3.3-28 pg/m3 at IADN stations and over Lake Superior (Cortes et al., 1998; Jantunen et al., 2000) (Table 2). HEPX

concentrations in Alabama air during May and August (21 and 54 pg/m3) were 2-8 times higher than those measured

over Lake Superior (2.8 and 12 pg/m3) (Jantunen et al., 2000) during the same two months. HEPT in Alabama air

averaged 26 pg/m3, but was below detection (<0.2 pg/m3) over Lake Superior in May and very low in August (0.8

pg/m3) (Jantunen et al., 2000). Thus HEPT exceeded HEPX in Alabama, whereas HEPX was the dominant compound

over Lake Superior.

Other OCs were similar in Alabama and the Great Lakes region. Concentrations of γ-HCH averaged 50 pg/m3

in Alabama, 19 pg/m3 over Lake Superior (Jantunen et al., 2000), 19 pg/m3 at Eagle Harbor and 29 pg/m3 at Sturgeon

Point (Table 2) (Cortes et al., 1998). Dieldrin averaged 38 pg/m3 in Alabama, 24 pg/m3 over Lake Superior (Jantunen et

al., 2000), 26 pg/m3 at Sturgeon Point and 11 pg/m3 at Eagle Harbor (Cortes et al., 1998). DDE in Alabama averaged

10 pg/m3 due to one high sample but was usually in the 1-6 pg/m3 range (Table 1). Concentrations of ΣDDT ranged

from 1.6-4.5 pg/m3 at Eagle Harbor and 5.5-34 pg/m3 at Sturgeon Point with the p,p’DDT/p,p’-DDE ratio ranging from

0.2-0.8 (Cortes et al., 1998).

Seasonal Trends and Relationship to Temperature:

OCs in air were generally higher in summer-fall than winter-spring. Exceptions were HEPT which was fairly

constant throughout the year and p,p’-DDE which also showed little seasonality except for a large spike in June (Figure

3). α-HCH was excluded due to breakthrough problems. The relationship to temperature was investigated by plotting

log P (partial pressure, Pa) versus 1/T (T = ambient air temperature, K), a version of the Clausius Clapeyron equation

(Cortes et al., 1998; Hoff et all., 1992). Results are shown in Figure 3. Regressions were carried out in two ways (a) by

including all the data points and b) by omitting one low-temperature point (1/T = 0.00365) and one outlying point at 1/T

= 0.0034. The reason for leaving out the low-temperature point is that in some cases the regressions are dominated by

this one value.

85

Table 3: Mean Atmospheric Concentrations (pg/m3) of OCs in the Great Lakes

Compound Alabama South 1 Eagle 2 Sturgeon 2 Lake 4

Carolina Harbor Point Superior

a-HCH 92 81 75 70g-HCH 50 19 29 19TC 47 100 4.3 10 6.8CC 25 50 3.9 11 4.7TN 23 30 2.8 6.8 3.8CN 3.1 0.9HEPT 26 1.2HEPX 16 7.4OXY 6.5 2.4p,p'-DDE 10 2.2 17Dieldrin 38 11 26 24

Toxaphene 176 189 6.55 3 20

1: Average concentrations for 1994-95 (Bidleman et al., 1998)2: Average concentrations for 1995-97 (Cortes et al., 1998)3: Average concentrations from Glassmeyer and Hites, 1999.4: Average concentrations for August 1996 and May 1997 (Jantunen et al., 1999)

Figure 3: Plots of Log P (partial pressure, Pa) versus 1/T (ambient air temperature, K). TC = trans-chlordane, CC = cis-chlordane, TN = trans-nonachlor, Tox = toxaphene, dieldrin and γ-HCH. The solid line is the linear regression using all data points, the dashed line is the linear regression line after 1-2 outlier points (shaded) are removed.

3.2 3.3 3.4 3.5 3.6 3.7-11

-10

-9

TN

3.2 3.3 3.4 3.5 3.6 3.7-11

-10

-9

-8

Tox

Log

P

1/T (K) x1000

3.2 3.3 3.4 3.5 3.6 3.7-11

-10

-9

-8

TC

3.2 3.3 3.4 3.5 3.6 3.7

-11

-10

-9

CC

3.2 3.3 3.4 3.5 3.6 3.7

-11

-10

-9

-8

Dieldrin

3.2 3.3 3.4 3.5 3.6 3.7-11

-10

-9

-8

g-HCH

86

The slope for toxaphene, using all the data, was -3350 with r2 = 0.39. Although the regression is significant

(p<0.05) the relationship is driven by one very low concentration in January (1/T = 0.00365). Removing this point

reduces the slope to -2065 and r2 = 0.14 (not significant, p>0.05) (Table 3 and Figure 3). The scatter of points and non-

significant regression implies that parameters other than temperature alone are controlling the ambient air concentration.

In particular, soil moisture is likely to be important for pesticide volatilization (Scholtz et al., 1997), and during the

summer months volatilization may be retarded as the soil dries out. Other influential factors are plowing events and

amount of vegetation cover. Advection of toxaphene from other areas of the southern U.S. may confound the local

temperature relationship. The poor correlation may also be a limitation of this small data set. Slopes found for

toxaphene in other recent studies are: Columbia, South Carolina (-2308) (Bidleman et al., 1998), Point Petre, Lake

Ontario (-2284) (Shoeib et al., 1999), Eagle Harbor, Lake Superior (-2438) (Glassmeyer and Hites, 1999) and Lake

Superior (-2500) (Jantunen et al., 2000).

The cyclodienes, including TC, CC, TN, CN, dieldrin and the metabolites HEPX and OXY, all showed the same

temporal pattern, lower levels in winter and early spring and higher levels in late spring, summer and early fall. HEPT

was the odd chlordane compound that did not show a seasonal trend. Lack of temperature dependence for HEPT was

found by Hoff et al., (1992) in southern Ontario although they found that the other chlordanes showed a significant

temperature dependence. Log P for TC, CC, TN, HEPX and OXY was significantly (p<0.01) correlated with 1/T, with

r2 values = 0.40 to 0.55 and slopes = -3468 to -3936 (Table 3 and Figure 3). Removing the two questionable points

(see above) had only a minor effect on the slopes, although the r2 values were improved (Table 3). As noted above,

release of chlordane from urban areas (cities of Muscle Shoals and Florence abut the TVA reservation) may be driving

the atmospheric levels rather than volatilization from agricultural soils and such release may be temperature-driven. The

slopes found here are similar to those found by Bidleman et al., (1998) in South Carolina for the sum of chlordanes

(TC+CC+TN) in the 1977-88 (-3626), but are steeper than the Columbia slopes in 1989-92 (-2129) and 1994-95 (-

2526). At the Great Lakes IADN stations, the slopes for the chlordane-related compounds ranged from -1674 to -3787

(Cortes et al., 1998).

Lindane showed a symmetric seasonal trend, peaking in June and dropping off faster in late summer than

toxaphene or the cyclodienes (Figure 3). The temperature slopes (r2 values) by including all points or excluding two

questionable points were -2614 (0.41) or -2081 (0.36). Regressions were significant (p<0.01) in both cases.

The temperature slopes for compounds with significant regressions are less than their vapor pressure slopes,

Table 4 (Hinckely et al., 1990). Slopes for Henry’s law constants are -2382 to -3005 for γ-HCH (Jantunen and

Bidleman, 2000; Kucklick et al., 1991) and -3209 for technical toxaphene (Jantunen and Bidleman, 2000). Hoff et al.

(1998) and Wania et al. (1998) argued that temperature slopes for compounds in ambient air might be lower than those

of their physicochemical properties if there is a substantial contribution from long-range transport compared to local air-

surface exchange. Apparent heats of air-surface exchange for OCs pesticides in the U.K. were similar to their heats of

vaporization when air samples were taken during stable atmospheric conditions; heats of exchange were lower during

unstable events (Lee et al., 2000). As mentioned above, other factors such as reduced water evaporation in dry periods

might retard volatilization in the hottest period of the year and thereby lower observed slopes.

87

Enantiomer Composition of Chiral Compounds

Chiral compounds determined in Alabama air included α-HCH, the chlordane components TC, CC and MC-5

(octachlordane) and metabolites HEPX and OXY. Although quantitative results for α-HCH were compromised by

breakthrough problems, determinations of enantiomer ratios (ER) were unaffected since breakthrough should be the

same for both enantiomers. The ER was calculated as the ratio of (+)/(-) enantiomers for α-HCH, TC, CC, HEPX and

OXY, and as the ratio of the first/second eluting peak for MC-5.

α-HCH was racemic in all air samples, with average ERs of 1.01±0.04 (BGB-172) and 0.99±0.02 (Beta-DEX).

The chlordane components TC, CC and MC-5 were nearly racemic, with ERs averaging 0.97±0.02, 1.00±0.04 and

0.97±0.03, respectively. The racemic values of chlordanes are consistent with

Table 4: Regression Parameters of log P versus 1/T Plots.subcooled liquid

all data 1-2 points removed vapr pressure (30)

compound slope intercept r 2 a slope intercept r2 a slope

g-HCH -3614±651 -0.56 0.41 -2081±590 -2.31 0.36 -3680TC -3468±880 2.11 0.45 -3622±802 2.70 0.56 -4216CC -3936±804 3.44 0.52 -4007±789 3.74 0.61 -4284TN -3772±719 2.81 0.55 -3848±960 3.14 0.78 -4468HEPT -767±810 -7.09 0.04 -3995HEPX -2720±703 -0.77 0.41 -2964±566 0.09 0.57OXY -3317±850 0.77 0.40 -3420±758 1.18 0.57Dieldrin -3653±825 2.75 0.46 -3491±851 2.24 0.57 -4310

Toxaphene -3350±873 2.05 0.39 -2065±1085 -2.06 0.14 -4487 b

a Values of r2 in bold are significant at p<0.01. b Average of two congeners (30).

evaporation from termiticide-treated houses. Racemic chlordanes were also found in the five indoor air samples from

Muscle Shoals. Average ERs for TC, CC and MC-5 = 0.98 ± 0.0095, 1.00 ± 0.012 and 1.00. The ER of MC-5

determined in one indoor air sample was 1.00 (Bidleman unpublished). Racemic ERs (1.00 ± 0.01, n=33) were also

found in a larger set of samples from the midwestern U.S. (mean ERs for TC and CC = 1.00 ± 0.01 (n=33)). By contrast,

soil in Alabama contained non-racemic chlordane residues. ERs for TC = 0.92±0.05, CC = 1.12±0.09, and MC-5 =

0.73±0.17 (Wiberg et al., 1997). These enantiomeric measurements and results from the soil-air exchange model

(Harner et al., 2000) indicate that volatilization from soils is unlikely to account for the chlordane levels in Alabama air.

Evaporation from termiticide-treated houses may be responsible for maintaining current atmospheric levels.

HEPX is produced mainly by the microbial degradation of HEPT in the soil and it is a minor product of HEPT

photolysis (Bidleman et al., 1998). The average ER of HEPX in Alabama air (1.77±0.20) was lower than for Alabama

soils (ER = 2.7-3.4) (Wiberg et al., 1997). However the level of HEPX in Alabama soils was low (geometric mean 0.1

ng/g dry weight) (Harner et al., 1999) and the ERs were determined for only three soil samples. Soils from Midwestern

states show great variability in the HEPX ERs ranging from 1.17-7.27 (Aigner et al., 1998).

OXY is a major metabolite of octa- and nonachlordanes and Buser and Müller (1992) reported ERs >1.00 in

biological samples. Reactions of (+) TC with CrO3 produces (+) OXY and (-) TC produces (-) OXY (Müller and Buser,

88

1994). The ER of OXY in Alabama air averaged 1.08±0.08. In soil, ERs of OXY ranged from 1.17-1.51 (n=4), but in

many samples the levels were too low for enantiomeric analysis. Midwestern U.S. soils show ambivalence in OXY ERs

(Aigner et al., 1998), ranging from 0.67-1.73. EFs of OXY in four British Columbia soils were 0.58-0.85 ( (Falconer et

al., 1997). Since chlordane residues in soil show depletion of (+) TC and (-) CC, the ER of OXY in soil is probably the

net result of (+) and (-) OXY production from chlordane metabolism.

Literature Cited Aigner, E., Leone, A., Falconer, R.L. 1998. Concentrations and enantiomeric ratios of organochlorine pesticides in soils from the U.S. cornbelt, Environ. Sci. Technol. 32, 1162-1168. Alegria, H.A., Bidleman, T.F., Shaw, T.J. 2000. Organochlorine pesticides in the ambient air of Belize, Central America. Environ. Sci. Technol. 34, 1953-1958. Andrew, P., Vetter, W. 1995. A systematic nomenclature system for toxaphene congeners Part 1: Chlorinated bornanes. Chemosphere 31, 3879-3886. Bidleman, T.F., Alegria, H., Ngabe, B., Green, C.1998. Trends of chlordane and toxaphene in ambient air of Columbia, South Carolina. Atmos. Environ. 32, 1849. Bidleman, T.F., Jantunen, L.M.M., Wiberg, K., Harner, T., Brice, K.A., Su, K., Falconer, R.L., Leone, A.D., Aigner, E.J., Parkhurst, W.J. 1998. Environ. Sci. Technol. 32, 1546-1548. Billings, W.N., Bidleman, T.F. 1980. Environ. Sci. Technol. 14, 679-683. Buser, H.-R., Hagland, P., Muller, M.D., Rappe, C. 1998. Organohalogen Compounds. 35, 239. Buser, H.R., Müller, M.D. 1995. Isomer and enantioselective degradation of hexachlorocyclohexane isomers in sewage sludge under anaerobic conditions. Environ. Sci. Technol. 29, 664-672. Cortes, D.R., Basu, I., Sweet, C.W., Brice, K.A., Hoff, R.W., Hites, R.A. 1998. Temporal trends in gas phase concentrations of chlorinated pesticides measured at the shores of the Great Lakes. Environ. Sci. Technol. 32, 1920-1927. Dearth, M.A., Hites, R.A. 1991, 25, 245-254. Complete analysis of technical chlordane using negative ionization mass spectrometry Environ. Sci. Technol. Falconer, R.L., Bidleman, T.F., Szeto, S.Y. 1997. Chiral pesticides in soils of the Fraser Valley, British Columbia. J. Agric. Food Chem. 45, 1946-1951. Finizio, A., Bidleman, T.F., Szeto, S.Y. 1998. Emission of chiral pesticides from an agricultural soil in the Fraser Valley, British Columbia. Chemosphere, 36, 345-355. Frenzen, G., Hainzl, D., Burhenne, J., Parlar, H. 1994. Structure elucidation of the three most important toxaphene congeners by X-ray analysis. Chemosphere 28, 2067-2074. Glassmeyer, S.T., Brice, K.A., Hites, R.A. 1999. Atmospheric concentrations of toxaphene on the coast of Lake Superior. J Great Lakes Res 25, 492-499. Hainzl, D., Burhenne, J., Parlar, H. 1994. Theoretical consideration of the structure variety in the toxaphene mixture taking into account recent experimental results. Chemosphere 28, 245-251. Halsall, C.J., Bailey, R., Stern, G.A., Barrie, L.A., Fellin, P., Muir, D.C.G., Rosenberg, B., Rovinsky, F.Ya, Kononov,

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E.Ya., Pastukhov, B. 1998. Environ. Pollut. 102, 51-62. Harner, T., Wideman, J.L., Jantunen, L.M.M., Bidleman T.F., Parkhurst, W.J. 1999. Residues of organochlorine pesticides in Alabama soils. Environ. Pollut. 106, 323-332. Harner, T., Bidleman, T.F., Mackay, D. 2001. Soil-air exchange model of persistent pesticides in the U.S. Cotton Belt. Environ. Toxicol. Chem. 20, 1612-1621. Hinckley, D.A., Bidleman, T.F., Foreman, W.T., Tuschall, J.R. 1990. Determination of vapor pressures for nonpolar and semipolar organocompounds from gas chromatographic retention data. J. Chem. Eng. Data, 35, 232-237. Hoff, R., Muir, D., Grift, N. 1992. The annual cycle of polychlorinated biphenyls and organochlorine pesticides in air in southern Ontario: 1 Air concentration data. Environ. Sci. Technol. 26, 266-275.. Hoff, R., Muir, D., Grift, N. 1992. The annual cycle of polychlorinated biphenyls and organochlorine pesticides in air in southern Ontario: 2. Atmospheric transport and sources. Environ. Sci. Technol. 26, 276-283. Hoff, R.M., Brice, K.A., Halsall, C.J., 1998. Nonlinearity in the slopes of Clausius-Claeyron plots for SVOCs. Environ. Sci. Technol. 32, 1793-1798. Iwata, H., Tanabe, S., Sakai, N., Tatsukawa, R. 1993. Distribution of persistent organochlorine pollutants in oceanic air and surface seawater and the role of ocean on their global transport and fate. Environ. Sci. Technol. 27, 1080-1098. Jantunen, L.M.M. Bidleman, T.F. 1998. Organochlorine pesticides and enantiomers of chiral pesticides in Arctic Ocean water. Arch. Environ. Contam. Toxicol. 35, 218-228. Jantunen, L.M., Bidleman, T.F. 2003. Air-water gas exchange of toxaphene in Lake Superior, Environ. Toxicol. Chem. 22, 1229-1237. Jantunen, L.M., Bidleman, T.F. 2001. Temperature dependent Henry’s law constant for technical toxaphene. Chemosphere, Global Change Sci. 2, 225-231. Keller, C.D., Bidleman, T.F. 1984. Collection of airborne polycyclic aromatic hydrocarbons and other organics with a glass fiber filter-polyurethane foam systems. Atmos. Environ. 18, 837-845. Klobes, U., Vetter, W., Luckas, B., Scherer, G. 1997. Organohalogen Comp. 31, 20-25. Lee, R.G., Burnett, V., Harner, T., Jones, K.C. 2000. Short term temperature dependent air-surface exchange and atmospheric concentrations of polychlorinated naphthalenes and organochlorine pesticides. Environ. Sci. Technol. 34, 393-398. Leone, A.D., Ulrich, E.M., Bodnar, C.E., Falconer, R.L., Hites, R.S. 2000. Organochlorine pesticide concentrations and enantiomer fractions for chlordane in Indoor air from the U.S. Cornbelt. Atmos. Environ. 34, 4131-4138. Müller, M., Buser, H.R. 1994. Identification and the (+)- and (-)-enantiomers of chiral chlordane compounds using chiral high performance liquid chromatography/chiroptical detection and chiral high-resolution gas chromatography/mass spectrometry. Anal. Chem. 66, 2155-2162. PHSU.S. Department of Health&HumanServices. Toxicological Profile for 4,4-DDT, 4,4-DDE, 4,4-DDD; Draft for public comment; U.S. Department of Health&HumanServices, Public Health Service, U.S. Government Printing Office: Washington, DC, 1992. Scholtz, M. T., McMillan, A. C., Slama, C., Li, Y.-F., Ting, N., Davidson, K. 1997. Pesticides Emission Modelling. Development of a North American Pesticides Emissions Inventory; Canadian Global Emissions Interpretation Centre, Ortech Environmental Canada: Mississauga, Ontario, 242 p.

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Saleh, M.A. 1991, Toxaphene chemistry, biology, toxicity and environmental fate. Reviews of Environmental Contamination and Toxicology. G.W. Ware. New York, Springer-Verlag, 118: 1-85. Shoeib, M., Brice, K.A., Hoff, R. 1999. Airborne concentrations of toxaphene congeners at Point Petre (Ontario) using gas chromatography-electron capture negative ion mass spectrometry (GC-ECNIMS). Chemosphere 5, 849-871. Shoeib, M., Brice, K.A., Hoff, R.M. 2000. Studies of toxaphene in technical standard and extracts of background air samples (Point Petre, Ontario) using multidimensional gas chromatography-electron capture detection (MDGC-ECD). Chemosphere 40, 201-211. Spencer, W.F., Singh, G., Taylor, C.D., LeMert, R.A., Cliath, M.M., Farmer, W.J., 1996, J. Environ. Quality, 25, 815-821. Stern, G.A., Muir, D.C.G., Ford, C.A., Grift, N.P., Dewailly, E., Bidleman, T.F., Walla, M.D. 1992. Isolation and identification of two major recalcitrant toxaphene congeners in aquatic biota. Environ. Sci. Technol. 26, 1838-1840. U.S. Department of Health & Human Services. Toxicological Profile for Aldrin and Dieldrin; TP-92/01; Public Health Service, U.S. Government Printing Office: Washington, DC, 1992. U.S. Department of Health & Human Services. Toxicological Profile for Chlordane; Draft for public comment; U.S. Department of Health & Human Services, Public Health Service, U.S. Government Printing Office: Washington, DC, 1992. U.S. Department of Health & Human Services. Toxicological Profile for Heptachlor and Heptachlor Epoxide; TP-92/01; U.S. Department of Health & Human Services, Public Health Service, U.S. Government Printing Office: Washington, DC, 1992. Vetter, W., Scherer, G. 1999. Persistency of toxaphene components in mammals that can be explained by molecular modeling. Environ. Sci. Technol. 33, 3458-3461. Voldner, E.C., Schroeder, W. H. 1989. Modelling of atmospheric transport and deposition of toxaphene into the Great Lakes ecosystem. Atmos. Environ. 23, 1949-1961. Walker, K., Vallero, D.A. Lewis, R.G. 1999. Factors influencing the distribution of lindane and other hexachlorocyclohexanes in the environment. Environ. Sci. Technol. 33, 4373-4378. Wania, F., Haugen, J.E., Lei, Y.D., Mackay, D. 1998. Temperature dependence of atmospheric concentrations of semivolatile organic compounds. Environ. Sci. Technol. 32, 1013-1021. Wiberg, K., Harner, T., Wideman, J.L., Bidleman, T.F. 2001. Chemosphere 45, 843-848..

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Paper 3

Air-Water Gas Exchange of Chiral and Achiral Organochlorine

Pesticides in the Great Lakes

Liisa M. Jantunen a*, Paul A. Helm b, Jeffrey J. Ridal c, Terry F. Bidleman a

2008. Atmospheric Environment, 42, 8533-8542.

a Centre for Atmospheric Research Experiments, Science and Technology Branch, Environment Canada, 6248 Eighth Line, Egbert, ON, L0L 1N0, Canada.

b Environmental Monitoring and Reporting Branch, Ontario Ministry of the Environment, 125 Resources Road, West Wing, Etobicoke, ON, M9P 3V6, Canada. c St. Lawrence River Institute of Environmental Sciences, Windmill Point, Cornwall, Ontario, K6H 4Z1, Canada. Contributions: Liisa Jantunen prepared, collected, processed and analysed samples from Lake Superior 1996 and

1997. Liisa and Jeff Ridal collected samples in July 1998 and Jeff collected samples in September 1998, they were

prepared, processed and analysed by Liisa. Jeff and Paul Helm collected samples from Lake Ontario in 2000, where

Paul prepared and processed samples; analysis was done by Liisa. Buoy samples were prepared, processed and

analysed by Liisa where Jeff setup and arranged the sampling. The manuscript was prepared by Jantunen with input

from the other authors. Terry Bidleman secured funding and provided scientific guidance throughout.

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Abstract

Parallel air and water samples were taken on lakes Superior, Huron, Erie and Ontario between 1996-2000 to

determine the occurrence and levels of hexachlorocyclohexanes (HCHs), trans-chlordane (TC), cis-chlordane (CC),

dieldrin (DIEL) and the metabolite heptachlor exo-epoxide (HEPX). Concentrations in the water varied greatly

among the lakes, but air concentrations were similar, thus resulting fugacity ratios varied by lake and compound.

The enantiomer fraction, EF = (+)/[(+) + (–)], was determined for α-HCH, TC, CC and HEPX. Enantioselective

degradation of (+)α-HCH was found in water of all the lakes where the depletion was greatest for Lake Superior,

which has the longest water residence time. Preferential loss of (+)TC and enrichment of (+)HEPX were found in

lakes Superior and Ontario, with similar EFs in both lakes. CC was racemic in Lake Superior and showed depletion

of the (+) enantiomer in Lake Ontario. Depletion of (+)α-HCH, (+)TC and enrichment of (+)HEPX was seen in all

air samples. CC varied from depletion of the (+) or (−) enantiomer and was racemic in some cases. Higher

concentrations of α-HCH in the air over Lake Superior were correlated with less racemic composition, providing

evidence of water-to-air exchange. Fugacity ratios for the HCHs approached equilibrium conditions within a factor

of 2, with slight excursions toward net volatilization or deposition, where generally higher excursions were seen for

the cyclodienes.

Introduction

For some time it has been known that the Great Lakes have undergone significant loadings of

organochlorine pesticides (OCPs) and that gas exchange is the dominant mechanism (Blanchard et al., 2004; Buehler

et al., 2000; Hillery et al., 1998; Hoff et al., 1996). Unlike wet and dry deposition, gas exchange is a reversible

process where deposition and volatilization occur simultaneously. Although, most OCPs were banned decades ago,

they are still found in the Great Lakes air (Buehler et al., 2000; Cortes and Hites, 2000; Gouin et al., 2005; Hoh and

Hites, 2004; Ridal et al., 1996; Sun et al., 2006a), precipitation (Simcik et al., 2000; Sun et al., 2006b) and water

(Blanchard et al., 2004; Buehler et al., 2000; L’Italien et al., 2000a,b; Marvin et al., 2004; Ridal et al., 1996; 1997).

Several OCPs are chiral, including α-HCH, components of technical chlordane, heptachlor (HEPT), o,p'-

DDT and some metabolites of these compounds. The parent chiral pesticides were manufactured as racemic

mixtures; i.e.; equal amounts of the (+) and (−) enantiomers. The enantiomer fraction (EF), defined as the amounts of

the (+)/[(+) + (–)] enantiomers, is 0.500 for the racemic pesticide and is altered by biological processes only. Abiotic

processes, such as transport phenomena, hydrolysis and photolysis, are not enantioselective. Enantioselective biotic

pathways include enzymatic degradation and preferential membrane permeation (Hegeman and Laane, 2002; Müller

and Kohler, 2004; Möller and Hühnerfuss 1993; Hühnerfuss et al. 1993; Ulrich et al., 2001).

Enantiomers are useful as marker compounds to follow transport processes. Freshly applied racemic pesticides

that undergo atmospheric transport from source regions remain racemic. Many OCPs have been banned in

industrialized countries and their legacy exists as residues in water bodies and soils, where they have undergone

partial degradation by microbial action. Enantiomer signatures aid in distinguishing microbially processed residues

that have been recycled from water and soil to the atmosphere (Bidleman and Falconer, 1999; Bidleman and Leone,

2004; Eitzer et al., 2003; Jantunen et al., 2008; Leone et al., 2001).

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Chiral signatures of OCPs other than α-HCH have not been reported for waters of the Great Lakes. These

are useful for tracing sources and air-surface exchange processes, as discussed above, and also for interpreting chiral

signatures of OCPs in the biota of the lakes. The purpose of this study was to determine the levels, net direction of

air-water gas exchange and EFs of some OCPs in lakes Superior, Erie, Huron and Ontario between 1996-2000.

Materials

All solvents were chromatographic quality (EM Science, Darmstadt, Germany). Sodium sulfate

(anhydrous, EM Science), alumina (neutral, 6% water deactivated, EMD Chemicals Inc. Gibbstown NJ, U.S.A.),

silicic acid (100 mesh size, 3% water deactivated, Mallinckrodt, Paris, KY, U.S.A.) and glass fiber filters (water:

GFF/F, 47 and 140 mm, nominal cut-off 0.7 µm; air: 20.3 x 25.4 cm, EPM 2000, collects 99% of particles larger

than 0.3 µm, Whatman, Maidstone, U.K.) were cleaned by baking at 400 oC. ENV+ cartridges (200 mg ENV+

Biotage, VA, U.S.A.) were cleaned by eluting with acetone. XAD-2 resin (Amberlite, macroreticular styrene

divinylbenzene copolymer, 20-60 mesh size, Supelco, Bellefonte PA, U.S.A.) and polyurethane foam plugs (PUF,

8.0 cm diameter x 7.5 cm or 6.8 cm diameter x 4.2 cm, Pacwill Environmental, Stoney Creek, ON, Canada) were

cleaned by Soxhlet extraction in acetone/dichloromethane and acetone/petroleum ether, respectively. The XAD-2

was packed into columns 15 cm i.d. containing 75 mL settled volume. Analytical standards used for quantification

were purchased from Supelco (Sigma-Aldrich, Bellefonte PA, U.S.A.) and labeled compounds were purchased from

Cambridge Isotope Laboratories (Andover, MA, U.S.A.).

Sampling Procedures

Air samples (Table 1) were collected over Lake Superior in August 1996 (n=4) and May 1997 (n=4); over

Lake Huron in both seasons (n=2, each) and over Lake Erie in August, 1996 (n=1). Air samples were also taken

over Lake Ontario and the St. Lawrence River in July (n=10) and September (n=4) 1998 and again in June 2000

(n=5) (Table 1). Air sampling was done by drawing ~400-1000 m3 air through a GFF followed by two PUF plugs

as described elsewhere (Jantunen and Bidleman, 2003). Four air samples were taken from a buoy in the west end of

Lake Ontario from June to September, 1998 (Table 1). These 350 m3 air samples were taken with a PS-1 sampler

(Tisch Environmental, Village of Cleves, OH, U.S.A.), consisting of a GFF (7.6 cm diameter) followed by one PUF

plug (6.8 cm diameter x 4.2 cm). The GFFs and PUFs were Soxhlet extracted in dichloromethane and petroleum

ether, respectively. Prior to extraction, the PUFs were spiked with surrogates to monitor recoveries: α-HCH-d6 was

added to 1996-1998 samples and in 2000 α-HCH-d6, γ-HCH-d6, 13C10-TN, 13C12-DIEL and 13C10-HEPX were added.

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Table 1. Concentrations of Gas Phase Organochlorine Pesticides in Air, pg m-3.

Lake αααα-HCH γγγγ-HCH TC CC TN CN HEPX DIEL

SuperiorAugust 1996

range 79-144 21-22 2.7-12 3.0-12 2.9-5.2 0.49-2.5 10-25 24-77mean 105 22 5.2 5.7 3.7 1.6 16 41s.d. 29 0.51 4.5 4.1 1.1 0.8 6.5 25n 4 4 4 4 4 4 4 4

May 1997range 36-66 9-28 2.4-30 1.5-13 1.5-10 0.3-0.4 1.9-3.3 7.7-17mean 50 16 6.1 3.1 3 0.4 2.8 11s.d. 12 7.1 2.5 0.9 0.8 0.04 0.6 3.8n 5 5 5 5 5 5 5 5

HuronAugust 1996

range 62-75 17-19 1.8-4.0 3.0-7.8 3.0-7.4 0.32-0.82 9 16n 2 2 2 2 2 2 1 1

May 1997range 37-41 20-27 2.4-30 1.5-13 1.5-10 0.8 4.1 27

n 2 2 2 2 2 1 1 1

ErieAugust 1996

63 48 20 21 21 1.8 40 122n 1 1 1 1 1 1 1 1

Ontario - shipJuly 1998

range 75-102 20-36 2.2-8.7 3.3-11 3.2-11 0.51-1.3 2.6-8.1 19-106mean 82 28 5.0 6.3 6.2 0.79 5.1 46s.d. 13 5.9 2.7 3.1 3.1 0.32 2.3 37n 10 10 10 10 10 10 10 10

September 1998range 32-47 12-20 3.2-8.2 3.5-10.4 3.2-9.4 0.4-1.0 2.5-9.91.12-3.9mean 37 16 5.6 6.6 6.0 0.7 6.2 2.5s.d. 7.1 3.9 2.1 3.0 2.6 0.2 5.2 2n 4 4 4 4 4 4 4 4

June 2000range 51-26 72-90 3.5-17 3.9-20 1.4-12 0.2-1.4 1.96mean 24 82 7.8 8.5 5.2 0.6s.d. 2.3 9 6.4 7.4 5.0 0.6n 5 5 5 5 5 5 1 0

Ontario - buoySummer 1998

range 22-39 4.4-13 5.3-17 2.5-9.1 1.4-12mean 31 9.1 12 6.2 5.2s.d. 8 3.5 5.3 3.2 5.0n 4 4 4 4 4 0 0

TC: trans-chlordane; CC: cis-chlordane; TN: trans-nonachlor; DIEL: dieldrin; HEPX: heptachlor exo-epoxide

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Low volume (LV, 4 L) surface water samples were collected in stainless steel cans for α- and γ-HCH in lakes

Superior (1996 and 1997, n=41), Huron (1996, n=6), Erie (1996, n=3) and Ontario (1998 and 2000, n=25) (Table

2). Depth profiles were also done at four stations on Lake Superior and at one station on Lake Ontario. Samples

were taken in GoFlo bottles on a rosette equipped with a conductivity temperature at depth (CTD). An α-HCH-d6

surrogate was added to the cans in 1996-1998, and in 2000 α-HCH-d6 and γ-HCH-d6 were added. HCHs were

concentrated by drawing water through a GFF, followed by a ENV+ cartridge using procedures described elsewhere

(Jantunen et al., 2004). Sampling cartridges were extracted with 12 mL acetone.

High volume (HV, 80L) water samples for β-HCH, trans- and cis-chlordane (TC, CC), trans- and cis-

nonachlor (TN, CN), heptachlor-exo-epoxide (HEPX) and DIEL were collected from Lake Superior (1996-1997,

n=25) and from Lake Ontario (1998, n=8; 2000, n=9) (Table 2). A previous paper reported air-water gas exchange

of toxaphene from the Lake Superior study (Jantunen and Bidleman, 2003). Sampling was done by collecting

surface water in stainless steel cans. Recovery surrogates were added, see air section. Additional recovery

experiments were done by spiking ~80 L of lake water with 4-5 ng of unlabeled cyclodienes (TC, CC, TN, CN,

DIEL) and metabolite (HEPX). The OCPs were concentrated by drawing water through a GFF and an XAD-2 resin

column. The resin was extracted with dichloromethane and dried over granular anhydrous sodium sulfate (Jantunen

and Bidleman, 2003; Jantunen et al., 2004).

Sample extracts were concentrated and exchanged into isooctane with a final volume of 1 mL. HV water

and air samples were cleaned up and fractionated on an alumina - silicic acid column, collecting two fractions (Helm

and Bidleman, 2003; Jantunen et al., 2000; 2003). LV water samples were cleaned up on an alumina column. If

needed, LV, HV and air sample extracts in isooctane were given a further clean up with 18 M sulfuric acid. Extracts

were reduced by nitrogen blowdown to ~100-1000 µL for quantitative and chiral analysis.

Analysis

Quantitative analysis was done on 30-m DB-5ms or 60-m DB-5 capillary columns (0.25 mm i.d., 0.25 µm

film thickness, J&W, Agilent Technologies, Palo Alto, CA, U.S.A.), using temperature programs similar to those

previously described (Jantunen et al., 2000). Instruments were a Hewlett Packard 5890 GC-5989B MS Engine or

Agilent 6890 GC-5973 MSD, both operated in the electron capture negative ion (ECNI) mode with methane reagent

gas (nominal pressure of 1.0 Torr for the MS-Engine and 2.2 mL min-1 for the MSD).

The ions monitored (target/qualifying) were: α-, β- and γ-HCH (255/257), α-HCH-d6 and γ-HCH-d6 (261),

DIEL (346/348 (MS-Engine) or 380/382 (MSD)), 13C12-DIEL (392), TC and CC (410/412), TN and CN (444/446),

13C10-TN (454), HEPX (386/388 (MS-Engine) or 316/318 (MSD)), 13C10-HEPX (328),

PCB-103 (326/328) and mirex (404). Sample volumes of 2 µL were injected splitless (split opened after 1.0 min).

Other conditions were: injector and transfer line temperature 250oC, ion source temperature

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Table 2. Concentrations of Dissolved Organochlorine Pesticides in Surface Water.

Lake αααα-HCH ββββ-HCH γγγγ-HCH TC CC TN CN HEPX DIEL

Superior ng L-1 pg L-1 ng L-1 pg L-1 pg L-1 pg L-1 pg L-1 pg L-1 pg L-1

August 1996, May 1997range 1.9-3.6 37-93 0.43-1.1 1.9-7.9 1.6-11 0.50-8.4 0.30-3.3 30-70 94-220mean 2.8 61 0.61 4.1 5.4 3.2 1.4 45 137s.d. 0.45 15 0.16 1.7 2.2 2.3 0.8 11 36n 41 25 41 25 25 25 25 25 25

HuronAugust 1996

range 1.1-1.2 0.93-0.96mean 1.1 0.94s.d. 0.096 0.013n 6 6

ErieAugust 1996

range 0.43-0.56 0.43-0.58mean 0.48 0.49s.d. 0.07 0.08n 3 3

OntarioJuly 1998

range 0.19-0.37 29-50 0.15-0.33 1.4-7.3 2.6-6.9 1.4-5.0 0.53-1.5 14-22 66-134mean 0.33 41 0.24 1.9 3.3 1.9 0.82 18 93s.d. 0.067 6.3 0.062 0.35 0.52 0.45 0.26 3.0 26n 25 8 25 8 8 8 8 8 8

June 2000range 0.19-0.37 27-53 0.28-0.33 1.9-5.0 2.6-5.8 1.0-2.2 0.25-0.40 7.1-13 62-97mean 0.26 42 0.30 3.2 4.2 1.5 0.29 10 83s.d. 0.079 14 0.022 1.7 1.8 0.66 0.07 2.1 14n 12 9 12 9 9 9 9 9 9

See Table 1 for cyclodiene abbreviations

150oC and quadrupole temperature 100oC. Quantification was carried out against five standards that spanned the

concentration range of the samples, using the Chemstation software. Mirex was added to extracts prior to injection

as the internal standard. Random samples were checked for native mirex and found negative in the Lake Superior

samples, sub-pg L-1 levels of mirex were found in fraction 1 of the Lake Ontario water samples, so PCB-103 was

used as the internal standard.

Several chiral columns were used for enantiomer analysis: β-DEX-120 (BDX, 20% permethylated β-

cyclodextrin in polydimethylsiloxane, 30 m x 0.25 mm i.d., 0.25 µm film thickness, Supelco, Bellefonte, PA,

U.S.A.), BGB-172 (BGB, 20% tert-butyldimethylsilylated β-cyclodextrin in OV-1701, 30 m x 0.25 mm i.d., 0.25

µm film thickness, BGB Analytik AG, Switzerland), and β-DEXcst (BDXcst, proprietary composition, 30 m x 0.25

mm i.d., 0.25 µm film thickness, Restek, Bellefonte, PA, U.S.A.). Different temperature programs were employed,

97

depending on the column and enantiomers separated; typical conditions are given in several references (Bidleman et

al., 2007; Bidleman et al., 2002, 2004; Jantunen and Bidleman, 1998; Shen et al., 2004, 2005; Wiberg et al., 2001).

The different columns were used to achieve the desired enantiomer separations, with confirmatory analysis on

columns with different elution characteristics where possible; e.g. α-HCH: BDX, BGB and BDXcst; TC: BDX and

BGB and CC: BDX, BGB and BDXcst; HEPX: BGB (Kurt-Karakus et al., 2005; Shen et al., 2005). Data presented

is an average of the primary and confirmation results.

Results and Discussion

Quality Control

Average recoveries of labeled surrogate chemicals were 72% for HCHs in LV and 68% in HV water

samples, and 62-108% for the other labeled surrogates in HV samples. Recoveries of unlabeled compounds from

HV samples ranged from 72-102%, after correcting for the native amounts in the water. Recoveries of labeled

surrogates in air samples ranged from 68-104%. Sample amounts were recovery corrected. HCHs were determined

in both LV and HV water samples, differences were not significant for any of the sampling years (p>0.05).

Blanks for α- and γ-HCH in LV samples were done by passing 100 mL deionized water through ENV+

cartridges. No peaks were seen; the lowest standard injected was 0.8 pg µL-1 yielding a method quantitation limit

(MQL) of ~100 pg L-1 (based on a 4-L water sample at 500 µL final extract volume). Blanks for XAD-2 were done

by passing 1-L deionized water through the column. Again, no peaks were seen and the MQL was estimated to be ~1

pg L-1 based on an 80-L sample and a final extract volume of 100 µL. Cans were rinsed with solvent between

samples. Air blanks were done by placing a filter and PUF plug in the sampling apparatus and drawing air for 30 s.

No peaks were seen and the MQL was 0.2 pg m-3 (2 pg m-3 for β-HCH, which has a lower ECNI-MS response)

based on a 500 m3 sample and a final extract volume of 100 µL. Concentrations of particulate OCPs from air and

water filters were also below MQL.

Air samples were taken in duplicate over Lake Ontario in July 1998 to determine the reproducibility of the

sampling and analytical techniques. For the five pairs, the percent difference was <10% for all compounds.

Racemic standards were repeatedly injected to determine the reproducibility of measuring EFs. Average EF

± standard deviation values were: α-HCH (0.499 ± 0.004, n=22), HEPX (0.502 ± 0.006, n=9), TC (0.498 ± 0.004,

n=19) and CC (0.501 ± 0.006, n=19) on the BGB column; α-HCH (0.500 ± 0.003, n=27) TC (0.501 ± 0.004, n=16),

CC (0.498 ± 0.003, n=16) on the BDX column and α-HCH (0.500 ± 0.001, n=23) and CC (0.499 ± 0.004, n=9) on

the BDXcst column. The criterion used for peak purity in samples was agreement of the target/qualifying ion ratio

within ± 5% of the standard values (Falconer et al. 1997).

Confirmation of α-HCH in air and water samples on BGB and BDX columns yielded an average EF

difference of 0.028% (n=132). EFs of TC on BDX and BGB differed by 0.3% (n=34); and EFs CC on the BGB and

BDXcst differed by 0.2% (n=20).

OCPs in Air and Water

HCHs in air

98

HCH concentrations in the air samples collected over the four lakes are reported in Table 1 and compared

to other studies in Table 3. Results for α-HCH are from shipboard samples only since breakthrough occurred on the

smaller PUF plugs used for sampling at the Lake Ontario buoy. Only α-HCH and γ-HCH were found in air; β-HCH

was always below the MQL. Air concentrations of α-HCH for an individual lake tended to be higher in the summer

months (July, August) than in spring (May, June) and fall (September). Considering these measurements and others

(Table 3), average concentrations of α-HCH across the basin spanned a fairly narrow range of 22-81 pg m-3, whereas

the range for γ-HCH was larger (9-165 pg m-3).

Lindane was registered for use in Canada until the end of 2003, mainly as a seed treatment and was

discontinued in the U.S.A. and Mexico in 2006. The heaviest use in Canada occurred on canola seed (300-500 t y-1) in

Saskatchewan, Alberta and Manitoba followed by corn seed (7-10 t y-1) in Ontario and Quebec (Li et al., 2004). High

concentrations of lindane in air have been reported in the Canadian prairies (Shen et al., 2004; Waite et al., 2001)

and an atmospheric transport model predicted that this region accounted for elevated air concentrations during the

summer of 1998 at Eagle Harbor, Lake Superior. This model under-predicted lindane concentrations in fall and

winter, when long-range transport from outside North America may have contributed to measured values (Ma et al.,

2003). The long-term variability in HCH concentrations measured at IADN stations on all five lakes was

investigated by Sun et al. (2006a). Concentrations of α-HCH in air declined between 1990 - 2001 with times for

50% decrease (t1/2) ranging from 3.1 - 4.1 y. Levels of γ-HCH declined more slowly, with t1/2 ranging from 4.2 - 10 y.

Seasonal trends in both compounds were related to temperature but γ-HCH also showed a spring application pulse,

referred to as a “spring peak” (Sun et al., 2006a).

In this study, levels of γ-HCH over lakes Superior and Huron were fairly uniform in August and May, and

for the most part did not show the "spring peak" (Buehler et al., 2001, 2002; Cortes et al., 1999; Garmouma and

Poissant, 2004; Sun et al., 2006a). Higher γ-HCH concentrations were found from shipboard measurements over

Lake Ontario in June, 2000 and over Lake Erie in August 1996: however, four air samples from the Lake Ontario

buoy between June - September, 1998 showed little variation in γ-HCH. The two July, 1998 samples from the buoy

averaged 31 pg m-3 for γ-HCH and compared well to shipboard Lake Ontario - upper St. Lawrence River air samples

collected in the same month, averaging 28 pg m-3. The lack of a well-defined spring peak for γ-HCH in this study

was probably because only a few air samples were collected in each month.

9

9

Table 3: Comparison to other Great lakes Air Measurements (pg m-3)

Location Year Season αααα-HCH γγγγ-HCH TC CC TN CN HEPX DIEL Reference

Great LakesSuperior

1996, 1997, summer 75 19 5.7 4.3 3.3 0.9 8.6 24 1May 2000-May 2001 81 23 2.1 1.9 0.7 nd 3.1 1.9 21996-2003 annual 79 18 2.7 3.3 2.3 nd 5.5 9.0 3July 2002 - June 2003 73 20 3.3 4.6 4.6 nd nd 18 4

Huron1996, 1997, summer 54 21 9.5 6.2 5.6 0.6 6.6 21 1May 2000-May 2001 37 21 1.2 1.0 3.6 nd 2.1 6.6 21996-2003 annual 26 9 1.9 2.4 2.4 nd 1.7 9.2 3July 2002 - June 2003 30 38 6.4 8.0 8.5 nd nd 28 4

Erie1996, 1997, summer 63 48 20 21 21 1.8 40 142 11996-2003 annual 84 31 10.0 11 7.7 nd 7.9 26 3July 2002 - June 2003 24 22 8.7 9.4 8.4 nd nd 50 4

Ontario1998, 2000, summer 48 37 6.9 8.2 6.2 0.82 5.0 37 1May 2000-May 2001 48 35 6.9 9.3 4.0 nd 8.5 12 21996-2003 annual 28 12 3.6 3.9 4.0 nd 2.9 13 3July 2002 - June 2003 22 32 20 19 15 nd nd 47 4July 2000 - June 2001 65 83 39 32 25 1.8 nd 51 5

Michigan1996-2003 annual 78 47 6.1 6.5 5.3 nd 10 24 31996-2003 annual, Chicago 52 35 46 39 23 28 110 3July 2002 - June 2003 29 26 41 38 21 nd nd 110 4

Southern Ontario May 2000-May 2001 58 81 4.4 3.1 0.87 nd 1.1 4.0 2(inland) July 2002 - June 2003 25 33 5.7 6.5 6.3 nd nd 54 4

July 2000 - June 2001 55 165 17 17 17 1.6 nd 72 5

nd: not determinedTC: trans-chlordane; CC: cis-chlordane; TN: trans-nonachlor; DIEL: dieldrin; HEPX: heptachlor epoxide

1: This Study, 2: Shen et al., 2004; 2005, 3: Sun et al., 2006a; 4: Gouin et al., 2005, 5: Harner et al., and Motelay et al., 2005.

100

Ratios of α-HCH/γ-HCH in air over North America are reflective of regional lindane use superimposed on a

background of technical HCH from long-range transport (Shen et al., 2004). A ratio of α-HCH/γ-HCH = 5-7 would

be expected from the composition reported by Iwata et al. (1993). Ratios of average concentrations (α-HCH/γ-

HCH) in air were higher over lakes Superior and Huron in August, 1996 (Superior = 4.8 and Huron = 3.8) than in

May, 1997 (Superior = 3.1 and Huron = 1.7). Ratios over lakes Erie and Ontario ranged from 0.3-2.9. As in this

study, Sun et al., (2006a) found higher ratios over Lake Superior than the other four lakes. Lower ratios in the lower

lakes reflect the local/regional usage of Lindane.

Passive air samplers were deployed across North America to determine annually integrated air

concentrations of HCHs (Shen et al., 2004). Higher ratios of α-HCH/γ-HCH in air were found on the east and west

coasts of Canada, in the Canadian Arctic, and at higher altitudes in the Canadian Rocky Mountains, while lower

ratios were found in the interior of southern Canada, the mid-Atlantic and southern U.S. states, Mexico and Central

America. Volatilization of α-HCH from Lake Superior surface water and from seawater along the coast of eastern

Canada raised the α-HCH/γ-HCH ratio in these regions (Shen et al., 2004). In the Great Lakes basin, Gouin et al.

(2005) found relatively low ratios in air (0.36-2.0) except on the shore of Lake Superior (3.7).

Cyclodiene pesticides in air

Cyclodiene compounds (TC, CC, TN, CN, HEPX, DIEL) in air are reported in Tables 1 and compared with

other reports in Table 3. The ΣCHLOR (TC+CC+TN+CN) concentrations over Lake Superior in August 1996 and

May 1997 showed little variation (average = 13-16 pg m-3). The ΣCHLOR over Lake Huron was similar in August,

1996 (14 pg m-3), but one high sample (54 pg m-3) raised the average to 30 pg m-3 in May, 1997. The single sample

collected over Lake Erie in August, 1996 contained ΣCHLOR = 64 pg m-3. The average ΣCHLOR over Lake

Ontario varied slightly from 18-28 pg m-3 in different seasons. Higher ΣCHLOR concentrations were sometimes, but

not always, found at the western end of the lake from shipboard and buoy sampling and in a July 1998 comparison of

samples collected simultaneously at the buoy and the Point Petre IADN station near the eastern end (Ridal et al.,

2000). Overall, the ΣCHLOR and individual species agreed well with other reports (Table 3). Elevated ΣCHLOR

have been reported in Toronto (Table 3) and Chicago, and are probably related to former termiticide usage (Gouin et

al., 2007; Motelay-Massai et al., 2005; Shen et al., 2005). Longer term studies at IADN stations identified a trend of

decreasing TC, CC and TN concentrations and t1/2 ranged from 6.1-13 y between 1996-2003, except for Brule River

on Lake Superior where there no significant trend for TC and TN (Sun et al., 2006a).

Ratios of average TC/CC and TC/TN concentrations in air tended to be lower in the summer-early fall

campaigns (0.54-0.91 and 0.56-1.5) than in spring (0.92-2.3 and 1.5-2.8) even though the ΣCHLOR varied little. At

IADN stations average TC/CC ratios ranged from 0.88-0.95 (Sun et al., 2006a). Gouin et al. (2007) examined 15

sites in the Great Lakes region using passive samplers deployed seasonally, and found TC/CC ratios generally in the

order summer ~ spring < fall ≤ winter. The lower proportion of TC in air during the summer is in accordance with

observations by Hoff et al. (1992a,b), who noted that the TC/CC ratio in southern Ontario air followed an annual

cycle, with a summertime minimum. Similar seasonal trends have been noted for arctic air (Bidleman et al., 2002;

Halsall et al., 1998; Hung et al., 2002; Oehme, 1991; Patton et al., 1991). Reasons for this are not known, but it has

101

been speculated that TC is more labile to photochemical reaction and therefore more readily lost in the season with

more sunlight (Halsall et al., 1998; Oehme, 1991; Patton et al., 1991). Another factor may be seasonally differing

contributions of chlordane sources to the atmosphere; e.g. termiticide and soil emissions, and volatilization from the

lakes. Higher ratios of TC/CC and TC/TN due to termiticide usage have been reported for indoor air (Leone et al.,

2000) and urban air in Toronto and Chicago (Gouin et al., 2007; Shen et al., 2005).

Higher concentrations of HEPX over Lake Superior were found in August, 1996 compared to May, 1997

(16 vs. 2.8 pg m-3), and the same was true for Lake Huron (9.0 vs. 4.1 pg m-3). HEPX concentrations over Lake

Ontario averaged 5.1 and 6.2 pg m-3 in July and September 1998. HEPX was detectable in only one of four air

samples taken in June 2000, at 2.0 pg m-3. For HEPX, the t1/2 ranged from 6.8-7.7 at IADN stations except Lake

Superior where there was no trend (Brule River) or an increasing trend (Eagle Harbor) (Sun et al., 2006a). DIEL

was detected in air samples in 1996-1998 at concentrations of 8-122 pg m-3 but analytical problems occurred in 2000

which prevented its determination. For DIEL, the t1/2 ranging from 4.9-9.6 at the IADN stations (Sun et al., 2006a).

HCHs in water

Concentrations of the three HCHs in the surface water of Lake Superior varied little at stations across the

lake, as indicated by relative standard deviations of 16%, 24% and 26% for α-HCH, β-HCH and γ-HCH. The

spatial variability of α-, β- and γ-HCH in Lake Ontario surface water in 1998 and 2000 was 20-30%, 16-33% and 7-

26% (Table 2). Mean concentrations of α-HCH in surface water increased: Ontario < Erie < Huron < Superior. A

different order was found for γ-HCH: Ontario < Erie < Superior < Huron, while β-HCH increased Ontario <

Superior (Table 2).

α-HCH and γ-HCH were determined in the deep water samples in Lakes Superior and Ontario. Slight, but

significant, trends were found of uniformly increasing concentrations of α-HCH (p < 0.001) and decreasing

concentrations of γ-HCH (p < 0.001) with depth in Lake Superior during August, 1996 (Figure 1a). Profiles in Lake

Ontario during June, 2000, showed that both HCHs decreased with depth with a sharp drop below ~20 m (Figure

1b). This is a contrast from the 1993 study of Ridal et al., (1996), in which higher α-HCH concentrations were

found in the hypolimnon. In 2000, situations where concentrations declined with depth correspond to net

depositional gas exchange (γ-HCH in Lake Superior, both HCHs in Lake Ontario), whereas lower concentrations of

α-HCH in the upper water column of Lake Superior correspond to net volatilization (see below). α-HCH entering

Lake Ontario from the Niagara River has a seasonal cycle thus leading to a seasonal cycle in Lake Ontario itself.

Highest concentrations in the river were found in the winter and early spring followed by declining concentrations in

the late summer (Williams et al., 2003). Williams et al. (2003) contributed this to air-water deposition in the colder

months and volatilization in the warmer months.

102

Figure 1: HCH concentrations with depth: a) Lake Superior, b) Lake Ontario.

Williams et al. (2001) and Marvin et al. (2004) reported concentrations (ng L-1 ) similar to ours of α-HCH

and γ-HCH in lakes Superior (1997, 2.8, 0.38), Erie (1998, 0.41, 0.32) and Ontario (1998, 0.40, 0.24). Levels of

HCHs in the Niagara River, which flows out of Lake Erie into Lake Ontario, have declined by 94% for α-HCH

between 1987-2000 (Williams et al., 2003) and γ-HCH dropped by 52-56% between 1986-1997 (Marvin et al.,

2004). Similar declines have been found in lake-wide measurements of Lake Ontario see Figure 2, although in

recent years γ-HCH has remained constant. If historical loadings of α-HCH can be assumed to result from rather

uniform air concentrations arising from long range transport from high usage regions such as India and China, higher

concentrations in lakes Superior and Huron can be explained by their longer water residence times and colder

average water temperatures compared to lakes Ontario and Erie. Water residence times (Botts and Krushelnicki,

1987) and annually averaged surface temperatures for 1992-2002 (from daily satellite data, NOAA) were: Superior

(191 y, 5.9oC), Huron (22 y, 8.5oC), Ontario (6 y, 9.6oC), and Erie (2.6 y, 11.3oC).

The α-HCH/γ-HCH ratio in surface water averaged 4.7 in Lake Superior, 1.2-1.3 in lakes Huron and

Ontario and 1.0 in Lake Erie. The lower ratios in lakes Huron, Erie and Ontario may reflect regional lindane usage.

The β-HCH/α-HCH ratios expected from the Iwata et al. (1993) technical HCH composition range from 0.07-0.20.

The proportion of β-HCH/α-HCH in Lake Ontario surface water (0.12) was within the range of the technical mixture

composition and higher than in Lake Superior (0.022), where β-HCH was depleted.

Lake Superior

0.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

0 100 200 300

Depth (m)

Con

cent

ratio

n (n

g/L)

αααα-HCH

γγγγ-HCH

a)

ng/L

Lake Ontario

0.15

0.20

0.25

0.30

0.35

0.40

0 50 100 150 200 250Depth (m)

Con

cent

ratio

n (n

g/L)

αααα-HCH

γγγγ-HCH

b)Lake Superior

0.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

0 100 200 300

Depth (m)

Con

cent

ratio

n (n

g/L)

αααα-HCH

γγγγ-HCH

a)Lake Superior

0.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

0 100 200 300

Depth (m)

Con

cent

ratio

n (n

g/L)

αααα-HCH

γγγγ-HCH

Lake Superior

0.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

0 100 200 300

Depth (m)

Con

cent

ratio

n (n

g/L)

αααα-HCH

γγγγ-HCH

αααα-HCH

γγγγ-HCH

a)

ng/L

Lake Ontario

0.15

0.20

0.25

0.30

0.35

0.40

0 50 100 150 200 250Depth (m)

Con

cent

ratio

n (n

g/L)

αααα-HCH

γγγγ-HCH

b)

ng/L

Lake Ontario

0.15

0.20

0.25

0.30

0.35

0.40

0 50 100 150 200 250Depth (m)

Con

cent

ratio

n (n

g/L)

αααα-HCH

γγγγ-HCH

ng/L

Lake Ontario

0.15

0.20

0.25

0.30

0.35

0.40

0 50 100 150 200 250Depth (m)

Con

cent

ratio

n (n

g/L)

αααα-HCH

γγγγ-HCH

b)

103

Figure 2: Decline of HCHs in Lake Ontario water (Ridal et al., 1996; Williams et al., 2003; L’Italien et al., 2000b).

Cyclodiene pesticides in surface water

Mean concentrations of ΣCHLOR, HEPX and DIEL in Lake Superior surface water in August 1996 and

May 1997 were 14, 45 and 137 pg L-1, respectively (Table 2). Lower concentrations were found in Lake Ontario in

1998 and 2000 where ΣCHLOR, HEPX and DIEL averaged 9.6, 14 and 88 pg L-1 (Table 2). Spatial differences

across the each individual lake were not found for any of the cyclodienes. Similar TC/CC and TC/TN ratios in lake

water were found in Lake Superior (0.71 and 1.2) and Lake Ontario (0.62 and 1.1).

Total chlordane concentrations have not changed at Niagara-on-the-Lake and Fort Erie between 1986-1998

(Williams et al., 2000), unlike DIEL and HEPX that have shown ~70% decline between 1986-2001 at the same

location (Williams et al., 2000; Marvin et al., 2004), reasons for this are unknown. Marvin et al. (2004) reported

DIEL water concentrations (pg L-1) in four Great Lakes (1997-2000) averaging: Superior 110, Huron 40, Erie 210

and Ontario 180. Williams et al. (2004) reported HEPX concentrations in Lake Superior ranging from 58-112 pg L-1

and <2-150 pg L-1in Lake Ontario.

Air-Water Gas Exchange

The net direction of gas exchange was determined from averaged air and water concentrations in each lake

and for the different sampling periods using the following fugacity relationships (Jantunen and Bidleman, 2003):

fW = CWH (1)

fA = CARTA (2)

FR = fW/fA = CWH/CARTA (3)

FF = fW/(fW + fA) = FR/(FR + 1) (4)

0

500

1000

1500

2000

2500

3000

3500

4000

4500

1986

1988

1990

1992

1994

1996

1998

2000

Year

ngL-1

0.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

4.0

4.5

5.0

α/γ

α/γ

α/γ

α/γ-

Rat

io

α-HCHγ-HCHα/γ ratio

0

500

1000

1500

2000

2500

3000

3500

4000

4500

1986

1988

1990

1992

1994

1996

1998

2000

Year

ngL-1

0.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

4.0

4.5

5.0

α/γ

α/γ

α/γ

α/γ-

Rat

io

α-HCHγ-HCHα/γ ratio

104

In equations 1-4, fW and fA are the fugacities in water and air (Pa), CW and CA are the dissolved concentration in water

and vapor-phase concentration in air (mol m-3), H is the Henry's law constant at the temperature of the water (Pa m3 mol-

1 ), R is the gas constant (8.314 Pa m3 mol-1 K-1), TA is the temperature of the air (K). FR is the water/air fugacity ratio

and FF is the fugacity fraction.

Henry’s Law constants were determined as functions of temperature by bubble stripping for α-HCH and γ-

HCH (Sahsuvar et al., 2003), TC, CC and TN (Jantunen and Bidleman, 2006) and DIEL and HEPX (Cetin et al.,

2006).

Values of FR <1.0 and >1.0 imply net deposition and volatilization respectively; and FR = 1.0 is air-water

equilibrium. Corresponding limits for FF are net deposition <0.5, net volatilization >0.5 and equilibrium = 0.5.

Uncertainties in fugacity calculations were estimated by propagation of errors in CA, CW and the Henry's law constants.

Uncertainties (relative standard deviations, RSD) for individual CA events were estimated as 0.1, based on observations

that CA for duplicate air samples taken over Lake Ontario varied by ~10% (Quality Control section). RSDs for CW were

based on the variability for water samples collected at different stations on a particular expedition (Table 2). RSDs for

Henry's law constants were 0.20 for α- and γ-HCHs (Sahsuvar et al., 2003), 0.33-0.30 for TC, CC and TN (Jantunen,

unpublished), and 0.26-0.29 for DIEL and HEPX (Cetin et al., 2006). RSDs in event FRs were propagated from the

above uncertainties (Sahsuvar et al., 2003) and the significance of net volatilization or deposition was judged from the

mean FR at p = 0.05.

Deposition and volatilization fluxes (ng m-2 d-1) were calculated from relationships presented in Jantunen and

Bidleman (2003), taking into account the variation in overall mass transfer coefficients (DAW, mol m-2 d-1 Pa-1) due to

wind speed (Table 4). The sign convention used here is negative for volatilization and positive for deposition. This is

consistent with the signage used in the IADN program (Blanchard et al., 2004) but opposite of that used in Jantunen and

Bidleman (2003). The net flux (Table 4) is the difference between deposition and volatilization.

Results for mean FF and FR (calculated from average FF according to eq 4) and net exchange direction for

individual events are summarized in Table 4 and significant events are described below. FRs for the HCHs approached

equilibrium conditions, generally within a factor of 2, with slight excursions toward net volatilization or deposition.

Volatilization of α-HCH was indicated from Lake Superior, May 1997 and Lake Huron, August 1996, and deposition in

Lake Ontario, July 1998. Deposition of γ-HCH was found in Lake Superior in both seasons and in Lake Ontario, July

1998 and June 2000. Volatilization of γ-HCH took place from Lake Huron, August 1996.

Net fluxes (ng m-2 d-1) for lakes Superior and Huron (August, 1996 and May, 1997) and Erie (August 1996)

ranged from -9 to -20 for α-HCH and -9 to 2 for γ-HCH. Those for Lake Ontario 1998 and 2000 were -3 to 13 for

α-HCH and 1.7 to 27 for γ-HCH.

Ridal et al. (1996) collected paired air and water samples from spring to fall, 1993 in Lake Ontario. They

found net volatilization of α-HCH during the summer months and net deposition in the spring and fall. As with this

study, γ-HCH in 1993 was generally undergoing deposition to Lake Ontario except for a few volatilization episodes

in late summer. Blanchard et al. (2004) at IADN stations predicted annual net volatilization of α-HCH from Lake

105

Superior and net deposition for the other four Great Lakes, whereas γ-HCH was undergoing net deposition in all

lakes, from 1995-2000.

Using the Jantunen and Bidleman (2006) Henry's law constants, FRs for TC, CC and TN were always >1

(range 1.8-6.4 for lakes Superior and Ontario, Table 4), indicating net volatilization. Volatilization was significant

for TC from Lake Superior in both seasons and Lake Ontario, July and September 1998; CC and TN from Lake

Superior, May 1997 and from Lake Ontario, July and September 1998. When lower "final adjusted value" (FAV)

Henry's law constants (Shen and Wania, 2005) were used, FRs of all chlordanes were not significantly different from

equilibrium (FRs = 0.63-1.4). Net fluxes of TC, CC and TN ranged from -2.3 to -0.5 ng m-2 d-1 for all the lakes using

Henry's law constants from Jantunen and Bidleman (2006) (Table 4), whereas TC and CC fluxes using the FAVs ranged

from -0.3 to 0.8 ng m-2 d-1.

Estimates of gas exchange for HEPX and DIEL were based on the Henry's law constants of Cetin et al.

(2006) or the FAV values of (Shen and Wania, 2005). Significant volatilization of HEPX occurred from Lake

Superior, May 1997 and Lake Ontario, July 1998. Deposition of DIEL was significant in Lake Superior, August

1996 and Lake Ontario, July 1998. For all sampling periods, HEPX fluxes ranged from -1.3 to 2.0 ng m-2 d-1 and

DIEL fluxes ranged from -8.1 to 0.6 ng m-2 d-1 (Table 4).

Enantiomer composition of chiral OCPs

α-HCH

Average EFs of α-HCH in air and water are summarized in Table 5 and shown in Figure 3. The EFs of α-

HCH in Lake Superior water did not vary between August, 1996 and May 1997, nor was there a significant trend

with depth (p = 0.2), but the EFs in air were closer to racemic in May than August. Similar trends were found for

Lake Huron. Lake Ontario water had the highest (closest to racemic) EFs of all the lakes sampled and unlike Lake

Superior, a significant trend was found for decreasing EFs from the surface (0.467) to 200 m (0.457) (r2 = 0.76, p

<0.05). EFs in air over Lake Ontario were generally <0.500 but higher than those in water, similar from ship and

buoy sampling and in all seasons (Table 5, Figure 3).

A study done by Ridal et al. (1997) on Lake Ontario during 1993 found an average EF in water of 0.462 ± 0.002.

EFs in air indicated nearly racemic α-HCH in spring and fall when the net gas exchange direction was depositional,

but the EFs were lowered in the warmer summer months when the net exchange direction was volatilization. In this

study, equilibrium or net deposition was estimated even in the summer months, and no seasonality in EFs in air were

found.

106

Table 4: Fugacity and Flux Calculations for the Great Lakes

Year Water Air Water Air Henry's

Temp Temp Conc Conc Law Constant mean FR mean FF Eventb K cOG Dd

AW Ue10 Flux (ng m-2 day-1)

(K) (K) pg L-1 pg m-3 (Pa m3 mol-1 K -1) fW fA-1 FR(FR-1)-1 decisions (m s-1) mol m-2 d-1 Pa-1 (m s-1) Deposition Volatilization Net

αααα-HCH a

LS-96 Aug 280 289 2760 105 0.12 1.30 0.57 VN 0.0042 0.151 4.8 38 -49 -11LS-97 May 275 278 2760 59 0.07 1.76 0.64 VY 0.0035 0.132 3.9 18 -27 -9.3LH-96 Aug 291 290 1140 69 0.31 2.30 0.70 VYY 0.0031 0.112 3.5 19 -39 -20LH-97 May 283 276 1140 39 0.15 1.96 0.66 VYN 0.0040 0.151 4.6 14 -26 -13LE-96 Aug 297 296 480 63 0.48 1.56 0.61 VN 0.0043 0.151 5.2 23 -35 -11LO-98 July 293 294 331 82 0.36 0.61 0.38 DY 0.0048 0.168 5.7 34 -20 13LO-98 Sept 291 291 331 37 0.31 1.13 0.53 VN 0.0067 0.237 7.9 21 -24 -3.0LO-00 June 285 289 264 24 0.19 0.85 0.46 DN 0.0044 0.157 5.0 8.9 -7.7 1.2

γγγγ-HCHLS-96 Aug 280 289 612 20 0.05 0.57 0.36 DY 0.0042 0.153 4.8 7.5 -4.5 2.9LS-97 May 275 278 612 16 0.03 0.52 0.34 DY 0.0036 0.133 3.9 4.9 -2.5 2.4LH-96 Aug 291 290 943 18 0.13 2.83 0.74 VYY 0.0032 0.115 3.5 5.0 -14 -9.2LH-97 May 283 276 943 23 0.06 1.13 0.53 VNN 0.0041 0.154 4.6 8.1 -9.2 -1.1LE-96 Aug 297 296 491 48 0.21 0.90 0.48 DN 0.0045 0.157 5.2 19 -16 2.4LO-98 July 293 294 240 28 0.16 0.56 0.36 DY 0.0049 0.173 5.7 12 -6.5 5.3LO-98 Sept 291 291 240 16 0.13 0.84 0.46 DN 0.0068 0.243 7.9 9.4 -7.7 1.7LO-00 June 285 289 303 82 0.08 0.12 0.11 DY 0.0044 0.158 5.0 31.0 -3.7 27

TCLS-96 Aug 280 289 4.1 5.2 14 4.95 0.83 VY 0.0013 0.048 4.8 0.6 -2.7 -2.1LS-97 May 275 278 4.1 6.1 11 3.33 0.77 VY 0.0012 0.046 3.9 0.6 -2.1 -1.5LO-98 July 293 294 1.9 5.0 24 3.98 0.80 VY 0.0011 0.038 5.7 0.5 -1.8 -1.3LO-98 Sept 291 291 1.9 5.6 22 3.18 0.76 VY 0.0016 0.059 7.9 0.8 -2.5 -1.7LO-00 June 285 289 3.2 7.8 17 2.58 0.72 VN 0.0012 0.043 5.0 0.8 -2.4 -1.6

CCLS-96 Aug 280 289 5.4 5.7 11 4.65 0.82 VN 0.0015 0.055 4.8 0.8 -3.3 -2.5LS-97 May 275 278 5.4 3.1 8.5 6.30 0.86 VY 0.0015 0.054 3.9 0.4 -2.5 -2.1LO-98 July 293 294 3.3 6.3 21 4.68 0.82 VY 0.0012 0.042 5.7 0.6 -3.0 -2.3LO-98 Sept 291 291 3.3 6.6 19 4.10 0.80 VY 0.0018 0.065 7.9 1.0 -4.2 -3.2LO-00 June 285 289 4.2 8.5 14 2.60 0.72 VN 0.0014 0.049 5.0 1.0 -2.9 -2.0

TNLS-96 Aug 280 289 3.2 3.7 11 4.10 0.80 VN 0.0015 0.054 4.8 0.5 -2.0 -1.5LS-97 May 275 278 3.2 3.0 8.4 3.83 0.79 VY 0.0015 0.055 3.9 0.4 -1.5 -1.1LO-98 July 293 294 1.9 6.2 24 3.18 0.76 VY 0.0011 0.038 5.7 0.6 -1.8 -1.2LO-98 Sept 291 291 1.9 6.0 22 2.94 0.75 VY 0.0017 0.060 7.9 0.9 -2.5 -1.6LO-00 June 285 289 1.5 5.2 15 1.79 0.64 VN 0.0013 0.046 5.0 0.6 -1.1 -0.5

HEPXLS-96 Aug 280 289 45 16.0 0.64 0.81 0.45 DN 0.0039 0.139 4.8 5.4 -4.0 1.3LS-97 May 275 278 45 2.8 0.46 3.20 0.76 VY 0.0033 0.124 3.9 0.8 -2.6 -1.8LO-98 July 293 294 18 5.1 1.48 2.23 0.69 VY 0.0041 0.145 5.7 1.8 -3.8 -2.0LO-98 Sept 291 291 18 6.2 1.31 1.81 0.64 VYN 0.0058 0.208 7.9 3.1 -4.9 -1.8LO-00 June 285 289 10 2.0 0.90 1.85 0.65 VN 0.0039 0.140 5.0 0.7 -1.3 -0.6

DIELLS-96 Aug 280 289 137 41 0.31 0.51 0.34 DY 0.0041 0.146 4.8 14.4 -6.3 8.1LS-97 May 275 278 137 11 0.22 1.20 0.55 VN 0.0035 0.129 3.9 3.3 -3.9 -0.6LO-98 July 293 294 84 46 0.79 0.86 0.46 DY 0.0045 0.158 5.7 17.8 -10.4 7.4

107

Table 4: Footnotes a) α- and γ-HCHs: hexachlorocyclohexanes; TC: trans-chlordane; CC: cis-chlordane; TN:

trans-nonachlor; DIEL: dieldrin; HEPX: heptachlor exo-epoxide b) V and D = net volatilization and deposition; N

and Y indicate significance for mean of individual FRs. The two events for Lake Huron, 1996-97 and Lake Ontario

Sept. 1998 have been coded individually. c) KOG: overall mass transfer coefficient, where kA (m s-1) = 0.001 +

0.000462 x [6.1 + (0.63 x U10)]0.5 x (U10 x Sc-0.67) (Mackay and Yuen, 1983). Sc is the gas-phase Schmidt

number, 2.58-3.25 for semi-volatile organochlorines (Bidleman and McConnell, 1995) d) DAW = 86400 s/d x KOG

/ RTA and e) U10: windspeed.

Figure 3: Enantiomer fractions of chiral OCPs in water (blue bar) and air (green bar).

Ena

ntio

mer

frac

tion

Water Air TC

SUP HUR ERIE ON-ship ON-buoy0.42

0.44

0.46

0.48

0.50

0.52

Aug. 96

May 97 A

ug. 96M

ay 97 Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

HEPX


0.60

0.62

0.64

0.66

0.68

0.70

0.72

Aug. 96

May 97

Aug. 96

May 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

α-HCH


0.44

0.46

0.48

0.50

0.52

Aug. 96

May 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98A

ug. 96M

ay 97

CC


0.48

0.50

0.52

0.54

0.56

Aug. 96

May 97

Aug. 96 M

ay 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

Ena

ntio

mer

frac

tion

Water AirWater Air TC


0.44

0.46

0.48

0.50

0.52

Aug. 96

May 97 A

ug. 96M

ay 97 Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

HEPX


0.60

0.62

0.64

0.66

0.68

0.70

0.72

Aug. 96

May 97

Aug. 96

May 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

α-HCH


0.44

0.46

0.48

0.50

0.52

Aug. 96

May 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98A

ug. 96M

ay 97

CC


0.48

0.50

0.52

0.54

0.56

Aug. 96

May 97

Aug. 96 M

ay 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

TC


0.44

0.46

0.48

0.50

0.52

Aug. 96

May 97 A

ug. 96M

ay 97 Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

TC


0.44

0.46

0.48

0.50

0.52

Aug. 96

May 97 A

ug. 96M

ay 97 Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

HEPX


0.60

0.62

0.64

0.66

0.68

0.70

0.72

Aug. 96

May 97

Aug. 96

May 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

HEPX

SUP HUR ERIE ON-ship ON-buoySUP HUR ERIE ON-ship ON-buoy0.58

0.60

0.62

0.64

0.66

0.68

0.70

0.72

0.58

0.60

0.62

0.64

0.66

0.68

0.70

0.72

Aug. 96

May 97

Aug. 96

May 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

Aug. 96

May 97

Aug. 96

May 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

α-HCH


0.44

0.46

0.48

0.50

0.52

Aug. 96

May 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98A

ug. 96M

ay 97

α-HCH


0.44

0.46

0.48

0.50

0.52

0.42

0.44

0.46

0.48

0.50

0.52

Aug. 96

May 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98A

ug. 96M

ay 97

CC


0.48

0.50

0.52

0.54

0.56

Aug. 96

May 97

Aug. 96 M

ay 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

CC


0.48

0.50

0.52

0.54

0.56

0.46

0.48

0.50

0.52

0.54

0.56

Aug. 96

May 97

Aug. 96 M

ay 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

Aug. 96

May 97

Aug. 96 M

ay 97

Aug. 96

Jul. 98S

ep. 98Jun. 00

Sum

mer 98

108

Table 5: Enantiomer Fractions of Chiral Organochlorine Pesticides in Surface Water and Air.

Lake αααα-HCH TC CC HEPX αααα-HCH TC CC HEPX

SuperiorAugust 1996

range 0.448-0.455 0.467-0.492 0.486-0.510 0.620-0.684 0.448-0.457 0.459-0.471 0.519-0.528 0.644-0.680mean 0.450 0.478 0.502 0.649 0.453 0.465 0.523 0.660s.d. 0.002 0.006 0.008 0.0182 0.004 0.005 0.004 0.016n 28 13 11 10 4 4 4 4

May 1997range 0.438-0.459 0.462-0.490 0.491-0.510 0.626-0.98 0.459-0.487 0.487-0.495 0.482-0.490 0.645-0.679mean 0.450 0.478 0.502 0.648 0.479 0.490 0.485 0.667s.d. 0.005 0.009 0.007 0.018 0.007 0.003 0.005 0.014n 37 12 9 12 5 5 3 5

HuronAugust 1996

range 0.454-0.464 0.462-0.471 0.462-0.465 0.515-0.517 0.669-0.672mean 0.459s.d. 0.005n 6 2 2 2 2

May 1997range 0.471-0.485 0.485-0.495 0.487-0.497 0.647

n 2 2 2 1

ErieAugust 1996

range 0.441-0.471 0.474 0.500 0.647 0.490 0.465 0.507 0.681mean 0.459s.d. 0.014n 3 1 1 1 1 1 1 1

Ontario - shipJuly 1998

range 0.468-0.479 0.452-0.488 0.468-0.510 0.623-0.701 0.481-0.496 0.466-0.485 0.496-0.510 0.639-0.682mean 0.469 0.477 0.480 0.655 0.491 0.476 0.502 0.663s.d. 0.003 0.009 0.033 0.027 0.004 0.005 0.005 0.018n 25 8 8 8 10 10 10 10

September 1998range 0.487-0.503 0.473-0.485 0.488-0.508 0.633-0.645mean 0.497 0.480 0.499 0.640s.d. 0.006 0.005 0.007 0.006n 3 4 4 4

June 2000range 0.462-0.476 0.472-0.494 0.483-0.495 0.647-0.670 0.485-0.497 0.467-0.485 0.504-0.524 0.624-0.681mean 0.461 0.483 0.490 0.658 0.491 0.478 0.516 0.656s.d. 0.001 0.006 0.006 0.010 0.006 0.007 0.007 0.029n 12 9 9 9 5 5 5 5

Ontario - buoySummer 1998

range 0.479-0.496 0.460-0.481 0.510-0.527 0.634mean 0.489 0.472 0.515s.d. 0.008 0.008 0.008n 4 3 3 1

TC: trans-chlordane; CC: cis-chlordane; HEPX: heptachlor epoxide

Water Air

109

Coupling between water and air is a function of the boundary layer stability (Honrath et al., 1997; Ridal et al., 1997).

The boundary layer over Lake Superior is stable in the summer months due to a strong inversion caused by the water

temperature being colder than the overlying air. The boundary layer becomes unstable when the water temperature

is warmer than the overlying air; this occurs in the winter months (Honrath et al., 1997). Based on lake area, the

boundary layer is probably most pronounced over Lake Superior, intermediate over Lake Huron and weakest over

lakes Ontario and Erie. Volatilization releases non-racemic α-HCH into the air boundary layer, where it mixes with

the α-HCH in background air. In August 1996, the EFs of α-HCH in the air and water of Lake Superior were very

closely coupled, differing by only 0.003. This implies that most of the α-HCH in the air above Lake Superior

during August 1996 was a result of revolatilization from the lake. The EFs of the air and water during May 1997

were not as closely coupled, differing by 0.026 (Table 5), implying that both advective and revolatilization processes

contributed α-HCH to the air over the lake. Application of a source apportionment model (Harner et al., 2000)

based on measured EFs in lake water and individual air sampled over the lake (at a height of ~10m) and an assumed

EF = 0.500 in background air (Shen et al., 2004), estimated that 90-94% of the α-HCH in the air during August 1996

originated from the lake. The volatilization percentage decreased to 42-66% in May. EFs in air and water of Lake

Huron were also less coupled in May than August, differing by 0.019 and 0.008. Estimates of the percentages of α-

HCH in air due to volatilization in these months were 80% in August and 51% in May. The α-HCH in air over lakes

Ontario and Erie were more decoupled from the water and showed the lowest contribution from the lakes (8-29%)

depending on the season. Higher air concentrations of α-HCH in air over lakes Superior and Huron were linearly

associated with less racemic EFs (Figure 4, r2 = 0.51 and p = 0.0057 for the combined data set). A similar plot for

Lake Ontario also suggested a decline in EFs associated with increasing air concentration, but the trend was not

significant (Figure 4, r2 = 0.06, p = 0.4). There are several reasons why Lake Ontario differed. Air-water exchange

of α-HCH in Lake Ontario was close to equilibrium or net deposition, whereas the other lakes showed net

volatilization. The boundary layer over Lake Ontario is likely to be smaller and less well developed because it is a

smaller lake. Finally, the EFs in Lake Ontario water were higher than in lakes Superior and Huron. Shen et al.

(2004) found that the enantiomer composition of α-HCH in ambient air of North America varied from depletion of

the (+) enantiomer along the east coast of Canada, nearly racemic inland, and depletion of the (−) enantiomer in

western Canada. Higher concentrations of α-HCH with EFs <0.500 were found on the north shore of Lake Superior.

Chlordanes and HEPX

Average EFs of these cyclodienes in air and water are summarized in Table 5 and Figure 3. TC in Lakes

Superior, Erie and Ontario was nonracemic, with average EFs ranging from 0.474-0.478. CC was racemic in Lakes

Superior and Erie (EFs 0.500-0.502), and nonracemic (EF = 0.480-0.485) in Lake Ontario. As for α-HCH, TC and

CC in air over lakes Superior and Huron tended to show greater deviation from racemic in August, 1996 than in

May, 1997 (Figure 3). EFs of TC in air over Lake Ontario were 0.472-0.480, similar to those in surface water, and

did not change seasonally. CC was racemic in air from ship measurements in summer 1998, but showed depletion of

the (–) enantiomer (EF = 0.515-0.516) in samples collected from the buoy during the same season and in ship

samples taken during June 2000.

110

Preferential degradation of (+)TC and (−)CC (EFs <0.500 and >0.500, respectively) was reported in

midwestern U.S.A. agricultural soils (Aigner et al., 1998) and overlying air (Bidleman and Falconer, 1999; Leone et

al., 2001). Similar enantiomer depletions have been reported for agricultural soils in Connecticut (Eitzer et al.,

2001). Chlordanes in air sampled away from cities at IADN stations (Ulrich and Hites, 1998), other sites in the

Great Lakes Basin (Gouin et al., 2007; Shen et al., 2005), the Arctic (Bidleman et al., 2002, 2004) and southern

U.S.A. (Vernier and Hites, 2007) showed preferential loss of (+)TC and (−)CC. Racemic chlordanes are associated

with residues from termiticide usage; e.g., in soil around house foundations that were treated with technical

chlordane (Eitzer et al., 2001). TC and CC in the indoor air of homes in the midwest U.S.A. were racemic (Leone et

al., 2000). The TC in ambient air samples from Muscle Shoals, AL (Jantunen et al., 2000; Shen et al., 2005),

Toronto (Gouin et al., 2007; Shen et al., 2005) and Chicago (Gouin et al., 2007) were closer to racemic than in rural

locations. Transport of chlordanes with different EF signatures to the lakes can be expected from different sources;

however, a lake vs. terrestrial source apportionment based on EFs, as for α-HCH, is difficult because the EFs of

chlordanes in background air are more variable and generally nonracemic.

For HEPX, enrichment of the (+) enantiomer was always found in air and water and the EFs were farther

from racemic compared to α-HCH, TC and CC. EFs of HEPX in water and air ranged from 0.620-0.701 and

0.624–0.682, respectively, with no significant differences for any lakes or years (Table 5, Figure 3). The enrichment

patterns of (+)HEPX in Great Lakes air and water were also found in air in earlier studies (Leone et al., 2001; Ulrich

and Hites, 1998) and in soils from the midwestern U.S.A. (Aigner et al., 1998; Leone et al., 2001), Alabama

(Jantunen et al., 2000; Wiberg et al., 2001), the Fraser Valley in British Columbia (Finizio et al., 1998; Falconer, et

al., 1997) and in air (Bidleman et al., 2002) and seawater (Jantunen and Bidleman, 1998) from the arctic.

Acknowledgments

We thank Janine Wideman and the crew/technical operators of the CCGS Limnos for help sampling, and Brian

Kerman, Environment Canada, for use of the Lake Ontario buoy.

111

Figure 4: Plot of αααα-HCH enantiomer fraction (EF) versus αααα-HCH air concentrations (pg m-3), dashed lines are the corresponding average EF of αααα-HCH in water.

Lake Superior

0.4450.4500.4550.4600.4650.4700.4750.4800.485

0 20 40 60 80 100 120 140 160

αααα-HCH Concentration (pg m -3)

αα αα-H

CH

EF

Lake Huron

0.460

0.465

0.470

0.475

0.480

0.485

0.490

0 10 20 30 40 50 60 70 80


αα αα-H

CH

EF

Lake Ontario

0.4650.4700.4750.4800.4850.4900.4950.5000.505

0 20 40 60 80 100 120


αα αα-H

CH

EF

Lake Superior

0.4450.4500.4550.4600.4650.4700.4750.4800.485

0 20 40 60 80 100 120 140 160


αα αα-H

CH

EF

Lake Huron

0.460

0.465

0.470

0.475

0.480

0.485

0.490

0 10 20 30 40 50 60 70 80


αα αα-H

CH

EF

Lake Ontario

0.4650.4700.4750.4800.4850.4900.4950.5000.505

0 20 40 60 80 100 120


αα αα-H

CH

EF

112

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Paper 4

Air-Water Gas Exchange of Toxaphene In Lake Superior


Environmental Toxicology and Chemistry, 2003, 1229-1237.

Environment Canada, 4905 Dufferin Street, Downsview, Ontario, M3H 5T4, Canada. Contributions: Liisa Jantunen prepared, collected, processed and analysed samples from both cruises on Lake Superior, carried out the data analysis and wrote the paper. Terry Bidleman secured funding and provided scientific input throughout the study.

118

Abstract

Parallel air and water samples were collected in Lake Superior, Great Lakes during August 1996 and May

1997, to determine the levels and air-water exchange direction of toxaphene. Concentration of toxaphene in water

did not vary across Lake Superior nor between seasons (averaging 918 ± 218 pg/L) but atmospheric levels were

lower in May (12 ± 4.6 pg/m3) than August (28 ± 10 pg/m3). Two recalcitrant congeners, Parlar 26 and 50, were

also determined. These congeners were enriched in the air samples, compared to a standard of technical toxaphene,

but not in the water. Water/air fugacity ratios varied from 1.4-2.6 in August and 1.3-4.7 in May, implying

volatilization of toxaphene from the lake. Net fluxes were estimated as 5.4-13 and 1.8-6.4 ng/m2d, respectively.

The temperature dependence of toxaphene partial pressure in air was Log P/Pa = -3291/T + 1.67. Using this

relationship, the atmospheric levels of toxaphene, fugacity ratios and net fluxes were estimated for the entire year.

Fugacity ratios were highest in the winter and lowest in the summer; thus toxaphene was predicted to undergo net

volatilization from the lake during all months. A yearly net removal of approximately 220 kg/y by gas exchange

was estimated.

Introduction

Semi-volatile organochlorine pesticides (OCPs) are atmospherically transported to the Great Lakes region

where they undergo air-water gas exchange, and to a lesser extent wet and dry deposition [1-3]. In the past, the

Great Lakes have been sinks for OCPs because they have large surface areas and are cold. Toxaphene is a complex

mixture of several hundred compounds [4], which was mainly used on cotton and soybeans in the southern U.S. and to

a lesser extent in other states; it also had limited use as a piscicide to eradicate rough fish. The peak usage of

toxaphene was 26-39 x 106 kg/yr between 1972-1975 [5]. Most registrations of toxaphene were cancelled in 1982, but

remaining stocks were applied through 1986 at a rate of 1-2 x 106 kg/y [5]. Evidence from peat cores [6] in the Great

Lakes region and western Canada, and sediment from Lake Ontario [7] and Siskiwit Lake on Isle Royale in Lake

Superior [8] indicate that atmospheric deposition of toxaphene and other OCPs peaked in the 1960s to 1970s and

declined into the 1980s. Sediments from Lake Superior itself show a varied picture, with peak accumulation years

ranging from 1974-1991 [9, 10]. Toxaphene residues remained constant with an insignificant rate of decline in

Lake Superior trout between 1977-1992, although levels declined in the other Great Lakes with half lives of 1.4-5

years [11, 12]. In 1999, 69% of Canadian fish consumption advisories for Lake Superior were due to toxaphene

[13]. Results from the Integrated Atmospheric Deposition Network (IADN) stations around the Great Lakes

indicate that between 1992 and 1996 atmospheric concentrations of several OCPs declined with a half time

decrease of 1.5-12.4 y [14]. However, time trends for toxaphene were not reported because it is not routinely

determined in the IADN program.

Swackhamer et al. [3] modeled the historical trend of toxaphene in the Great Lake atmosphere, using the

1989 atmospheric measurements in southern Ontario, Canada of Hoff et al. [15] and assuming that earlier air

concentrations were proportional to toxaphene production. Based on this trend, the estimated net air-water gas

exchange direction to Lake Superior was depositional throughout the 1970s and 1980s, but the net flux direction

was reversed to volatilization in the 1990s as atmospheric concentrations decreased. A budget for the mid-1990s

119

indicated that net gas exchange accounted for 70% of annual toxaphene losses from Lake Superior [3]. James et

al. [10] calculated toxaphene fluxes from air and water measurements during 1997-1998 and estimated that

toxaphene was volatilizing from Lakes Michigan, USA and Superior during 1997-1998. They found that Lake

Superior was 200-1000% saturated during the spring and summer months. Similar flux reversals have been

estimated for other OCPs [1,2]. Thus further declines in levels of OCPs in the Great Lakes waters are dependent on

reducing atmospheric levels [16].

This study evaluates the gas exchange of toxaphene using parallel air and water data collected in August

1996 and May 1997. Fluxes of total toxaphene are calculated for these months and estimated for a full year,

representative of the mid to late 1990s, based on the relationship between air concentration and temperature.

Profiles of individual congeners in air and water are compared and two persistent congeners are quantified.

Figure 1:Cruise track on Lake Superior showing station numbers (Table 1). Experimental Section

Cruise Track

The Canadian Coast Guard Ship Limnos left Sarnia, Canada on August 3rd 1996 and proceeded north

through Lake Huron to Sault St. Marie, Canada, at the mouth of Lake Superior. The ship followed the northern

shore of Lake Superior to Duluth, USA, and then the southern shore back to Sault St. Marie, Canada. The ship

then transversed Lake Huron and on August 18th arrived at Port Colborne, Canada in Lake Erie. The following

spring another sampling trip took place from May 12-27 1997. The same cruise track was followed, but leaving

23

51

68

80

95113

127

136139

Thunder Bay

157

169

196

201

221Duluth

Sault St.Marie2

120

and returning to Sarnia and not covering Lake Erie, see Figure 1 for cruise track and station numbering on Lake

Superior.

Water and Air Sampling

In August 1996 and May 1997, 14 and 11 surface water samples were collected on Lake Superior for

toxaphene. Sample details, location and hydrographic information are in Table 1. Samples of 80 L were taken via

a submersible pump. The water was passed through a glass fiber filter and the toxaphene was adsorbed on XAD-2

resin.

Four air samples were taken over Lake Superior, two over Lake Huron and one over Lake Erie in August

1996. In May 1997, five samples were taken over Lake Superior and two over Lake Huron. See Table 2 and

Figure 1 for sampling locations, atmospheric conditions and concentration data. Air sampling for toxaphene was

done by drawing 500-700 m3 of air, at a flow rate of 0.5 m3/min, through a glass fiber filter (Whatman, Maidstone,

England, 20.3 x 25.4 cm, EPM 2000, collects 99% of particles >0.3µm) followed by two plugs of polyurethane

foam (PUF), each 8 cm diameter x 7.5 cm tall. Methods for filter and absorbent preparation and sample extraction

have been published for water [17,18] and air [19].

Sample Preparation and Analysis

Cleanup and fractionation of air and water sample extracts were carried out on alumina-silicic acid columns, as

previously described [19]. Polychlorinated biphenyls, p,p’-dichlorodiphenylethane (p,p’-DDE), hexachlorobenzene

and mirex elute in fraction 1 (petroleum ether); toxaphene and most other OCPs elute in fraction 2 (dichloromethane),

although Parlar congener 26 (see below) splits between the two fractions. Toxaphene was determined by gas

chromatography electron capture negative ionization mass spectrometry on a Hewlett Packard 5890 GC-5989B MS

Engine with methane at a nominal pressure of 1.0 Torr. The column used for analysis was a DB-5 (J&W, 60 m x 0.25

mm i.d., 0.25 µm film thickness) operated at a helium carrier gas flow of 40 cm/s. Other conditions were as described

previously [19]. Toxaphene homologs sought were 7-Cl, 8-Cl and 9-Cl, monitoring ions (m/z values) of 343/345,

379/381 and 413/415 [18,19]. Mirex was added to extracts prior to injection as the internal standard. Both fractions

were checked for native mirex and found negative.

Toxaphene was quantified in all samples by the single response factor method (SRF) versus a technical

toxaphene standard obtained from the U.S. Environmental Protection Agency Repository for Pesticides and

Industrial Chemicals. A multiple response factor (MRF) method [17] was also used for the May set of air samples.

The two methods were compared for the May air samples because response factors of individual toxaphene

congeners by NIMS vary, and the MRF approach takes this into account. Two single toxaphene congeners were

quantified versus pure standards, see below.

121

Table 1: Water concentrations of dissolved toxaphene in Lake Superior, Great Lakes, pg/L

Station Water Total Parlar

Year Number Temperature oC Toxaphene P26 P50

1996 2 15.8 525 2.0 inter

1997 23 1.7 1196 9.2 23

1996 31 12.5 711 1.0 * 5.8

1997 31 2.5 969 0.4 * 15

1996 51 10.6 828 2.2 10

1996 68 15.0 685 2.0 7.6

1997 68 2.9 821 4.3 14

1996 80 3.6 1440 3.4 18

1997 80 2.2 1074 6.0 22

1996 95 4.0 693 inter 7.8

1997 95 1.8 1064 5.0 20

1996 113 4.4 1286 3.5 11

1997 113 1.9 1077 5.3 20

1996 127 5.2 1129 2.4 7.9

1996 136 15.6 859 inter 5.9

1996 139 15.6 810 1.3 * 4.5

1997 139 3.6 705 2.7 13

1997 157 1.8 982 4.7 18

1996 169 10.5 1127 3.2 15

1997 169 1.9 951 5.9 17

1996 196 10.5 743 1.8 * 4.0

1997 196 4.2 814 6.1 16

1996 201 14.2 848 1.3 * 7.0

1997 201 2.1 1089 5.6 24

1996 221 13.0 755 inter 7.9

1997 300 2.8 693 2.2 14

Mean 918 3.5 13

Standard Deviation 218 2.2 6.0

inter: interference or did not meet +20% ion ratio criteria

*below lowest quantitation standard.

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Table 2: Atmospheric Concentrations of Toxaphene over the Great Lakes, August 1996 and May 1997, pg/m3, see Figure 1 for sample locations.

Air Wind Total Parlar

Sample Location, Great Lakes Temperature, oC Speed, m/s Toxaphene P26 P50

1996

1 Lake Huron Sarnia, ON Canada - 19 3.3 29 3.0 1.1

Sault St Marie (SSM), ON Canada

2 Lake Superior SSM - station (st) 80 13 5.3 28 2.3 2.6

3 Lake Superior st 80 - st 139 9.3 5.7 17 1.9 1.7

4 Lake Superior st 139 - st 169 15 4.1 24 2.5 1.9

5 Lake Superior st 169 - st 51 17 4.4 41 2.1 1.3

6 Lake Huron st 51 - Sarnia 16 3.8 29 1.3 0.8

7 Lake Erie Sarnia - Port 23 5.2 39 4.3 1.4

Colbourne, ON Canada

Lake Superior Mean 28 2.2 1.9

Standard Deviation 10 0.3 0.6

1997

8 Lake Huron Sarnia - SSM 3.0 5.1 8.9 0.3 0.2

9 Lake Superior SSM - Thunder Bay, ON Canada 2.0 4.4 4.9 0.2 0.2

10 Lake Superior Thunder Bay - Duluth, MN USA 3.1 3.6 11 0.4 0.2

11 Lake Superior Duluth - st 157 3.8 4.2 12 0.3 0.3

12 Lake Superior st 157 - Thunder Bay 3.7 5.0 17 0.4 0.3

13 Lake Superior Thunder Bay - SSM 4.8 2.3 15 0.3 0.3

14 Lake Huron SSM - Sarnia 5.6 4.1 7.8 1.2 0.2

Lake Superior Mean 12 0.32 0.26

Standard Deviation 4.6 0.08 0.05

Quality Control

A surrogate of deuterated α-hexachlorocyclohexane (δ6 α-HCH) was added to each water sample and

yielded recoveries of 71 ± 21% (n=25). Additional recovery experiments were done by spiking approximately 80

L of lake water with 1150 ng toxaphene, yielding an average recovery of toxaphene 84 ± 9% (n=5), after

correcting for the native amounts in the water. Water sample concentrations were adjusted based on the average

toxaphene recovery. Blanks were done by passing 1 L of deionized water through an XAD-2 column and

analyzing as for a water sample, but no peaks were seen above the noise level (signal to noise was <2:1). The lowest

concentration of technical toxaphene standard injected was 12 pg/µL, signal to noise is approximately 10:1, giving a

method quantitation limit of approximately 75 pg/L, based on an 80-L water volume and a 500-µL extract volume.

123

Air spikes were done by fortifying clean PUF plugs with toxaphene (1150 ng), yielding average recoveries of 96 ±

3% (n=4). Air blanks were done by placing a clean filter and PUF in the sampling apparatus and drawing air for 30

s. As for water blanks, no peaks were observed above the noise level (signal to noise <2:1), and the method

quantitation limit was estimated at 3.4 pg/m3 based on a 500 m3 air sample and a 150 µL extract volume.


Toxaphene in Water

Results for water samples are given in Table 1. Average concentrations of toxaphene in water in August

and May were 848 ± 260 pg/L and 952 ± 161 pg/L. There is no statistical difference between the water data from

1996 and 1997 (p >0.2), so all samples were averaged, giving 918±218 pg/L. All water filters were analyzed but

no peaks were seen above the noise level.

Peaks matching in retention time and ion ratios (± 20%) to two persistent congeners were quantified against

pure standards (Parlar 5-component mixture, Axact Standards Commack, NY, USA). The two congeners were 2-

endo,3-exo,5-endo,6-exo,8,8,10,10 octachlorobornane and 2-endo,3-exo,5-endo,6-exo,8,8,9,10,10-

nonachlorobornane. The octachloro compound is also known as Parlar 26 (P26) [20], T2 [21] and B8-1413 [22] and

the nonachloro compound as P50, T12, B9-1679 and Toxicant Ac [23]. (The endo,exo nomenclature usage here is

consistent with (International Union of Pure and Applied Chemistry (IUPAC) rules [22],and differs from some reports

[20,21]). Although peaks matching P26 and P50 were quantified as these compounds, they may not be single

compounds in air or water samples. Shoeib et al. [24] showed this by using multidimensional gas chromatography-

electron capture detection (ECD) to examine the composition of toxaphene peaks in air samples collected on the north

shore of Lake Ontario. Concentrations of P26 and P50 in August and May ranged from 0.4 - 9.2 pg/L and 4.0 - 24

pg/L, respectively. The lowest standard injected gave a quantitation limit of 2 pg/L for both P26 and P50 and the

detection limit based on an estimated 2:1 signal to noise was 0.2 pg/L, results below the quantitation limit are flagged

in Table 1. The amounts found correspond to 0.40 ± 0.14 % and 1.4 ± 0.37% of total toxaphene. Reported

percentages of P26 and P50 in technical toxaphene are 0.52 % and 1.2 % (this work), 0.49 % and 1.5 % [24] and

0.40 % and 1.0 % [25]. The averages of these are 0.47 ± 0.062 and 1.2 ± 0.25. No significant differences between

the water and the technical standard were found (p >0.1).

Swackhamer et al. [3] reported an average total toxaphene concentration of 1120 ± 180 pg/L in Lake

Superior in 1994-96, also using XAD extraction and GC-NIMS. Their levels are about 20% higher than ours,

which is good agreement considering the difficulties associated with the determination of toxaphene (see below).

James et al. [10] reported a similar concentration found in this study of 900 ± 150 pg/L. Concentrations of

toxaphene reported by Swackhamer et al. [3] for other lakes were lower (pg/L): Huron = 470 ± 250, Michigan =

380 ± 120, Ontario = 170 ± 70 and Lake Erie = 230 ± 7. James et al. [10] also reported levels for Lake Michigan

of 407 ± 153 pg/L. During our study one sample was taken in Lake Erie and yielded 96 pg/L.

124

Toxaphene in Air

Toxaphene concentrations reported in Table 2 are for the gaseous phase only; all filters were analyzed and

toxaphene was below detection limits (<2:1 signal to noise). Toxaphene levels in air were higher in August than

May, averaging 28 ± 10 pg/m3 and 12 ± 4.6 pg/m3 using the SRF (Tables 2 and 3). For May, the average using the

MRF was 12 ± 3.1 pg/m3 and not significantly different from the mean based on the SRF. In a study of toxaphene

in Arctic Ocean water, similar agreement between the MRF and SRF was found [19]. In a survey of toxaphene in

Alabama air, where concentrations are an order of magnitude higher than in the Great Lakes region, the MRF

yielded results that averaged 20% lower than the SRF [19]. Results of these three comparisons suggest that the

SRF is not markedly different from the MRF in quantifying total toxaphene in air and water samples, perhaps

because differences in the individual response factors tend to cancel out when dealing with a large number of

congeners.

Concentrations of individual congeners in the air samples (mean + s.d., pg/m3) were: P26 = 2.2 ± 0.3 and 0.32 ±

0.08 and P50 = 1.9 ± 0.6 and 0.26 ± 0.05 for August and May. These correspond to percentages of total toxaphene

in air of 8.9 ± 2.7 % and 6.7 ± 2.8 % for August and 2.6 ± 4.5 % and 2.1 ± 0.7 % for May (percentages in technical

toxaphene are listed above). Both P26 and P50 were enriched in the air samples compared to the technical mixture

(p <0.01) for both seasons. Shoeib et al., [24] found similar percentages of total toxaphene in air at Point Petre,

Lake Ontario for P26 (mean = 7.1 %) and P50 (mean = 6.1%). Alabama air sampled in 1996-1997 showed

significant (p <0.05) enrichment of these two congeners (P26 = 1.9±2.0% and P50 = 3.3±3.4%) [19], but less than

for Lake Superior air. This implies preferential transport of P26 and P50 due to longer degradation times

compared to other congeners. Vetter and Scherer [26] showed that P26 and P50 have very stable structures

compared to other polychlorinated bornanes, due to the placement and orientation of the chlorine atoms on the

carbon backbone.

Historical measurements of atmospheric toxaphene around the Great Lakes are limited. McConnell et al.

[27] determined total toxaphene in air at Green Bay, WI in February, 1988 (15-26 pg/m3, n=2) and June, 1989 (30-

89 pg/m3, n=6). Toxaphene in four samples collected over lakes Michigan, Huron, Erie and Ontario in August,

1991 averaged 33 ± 24 pg/m3 [27]. The 1988-89 mean toxaphene concentration at Egbert, Ontario was 26 ± 32

pg/m3 [15]. Toxaphene in air at Eagle Harbor, Lake Superior, 1996-1997, averaged 4.8 ± 4.4 pg/m3 with one high

sample of 63 pg/m3 [28]. Toxaphene averaged 3.8 pg/m3 during 1995-1997 at Point Petre, Lake Ontario [24].

James et al. [10] found levels of total toxaphene above lakes Superior and Michigan in the spring and summer of

1997-1998 of 3-54 and 3-57 pg/m3, respectively. James et al. [10] also reported levels of total toxaphene at

Sleeping Bear Dunes on the shore of Lake Michigan of 19-70 pg/m3. With the exception of the Egbert [15] study,

all of these measurements were done by GC-NIMS.

The levels reported at Eagle Harbor (EH) [28] and Point Petre (PP)[15] are lower than levels reported

here, and there are several possible explanations for this; first being the EH and PP data represent the entire year,

whereas only spring-summer values are presented in this study. Secondly, the EH and PP were taken at land-based

stations; our samples were taken over water. Fugacity ratios (discussed later) predict toxaphene volatilization from

125

the water to the air, possibly increasing the concentration of toxaphene in the boundary layer over the lake. Lastly,

there may be analytical differences among the three laboratories. A recent inter-laboratory comparison of total

toxaphene in air showed a relative standard deviation of 55%, with a range of 47-84% from the mean value 29].

Much higher total toxaphene levels were found in Alabama, 1996-1997 (6 - 661, mean = 176 pg/m3) [19]

and South Carolina, 1994-1995 (39 - 483, mean = 189 pg/m3) [30]. These are past usage regions where toxaphene

was heavily applied to cotton and soybeans [5]. Concentrations in Belize, Central American (14-74, mean = 28

pg/m3) [31], were similar to Great Lakes values. In the Canadian Arctic, levels are lower or comparable to those in

the Great Lakes regions ranging from 3-7 pg/m3 by GC-NIMS [18,32].

Chromatographic Patterns of Toxaphene

Chromatograms of toxaphene are shown in Figure 2, where Cl-7 to Cl-9 homologs and total toxaphene

profiles in air and water are compared to those of the technical standard. The water and air have similar patterns,

both dominated by the lighter earlier eluting congeners and showing transformations from the standard mixture.

Toxaphene peaks are rarely made up of pure components, so the peaks have been numbered but known

components of the peaks have been identified [24, 33]. From Figure 2 in the Cl-8 homolog, Peak#1 contains P26

(see description above), Peak#2 contains a persistent octachlorobornane (no Parlar number, B8-1412) [22] that was

identified by using a standard supplied by Walter Vetter, University of Jena, Germany. Peak#3 contains P39 (B8-

531), Peak#4 P40 + P41 (B8-1414 + B8-1945), Peak #5 P42 (B8-806/809) and Peak#6 P44 (B8-2229) [33].

Peaks identified in the Cl-9 homolog are Peak#7 contains P50 (see description above) and Peak#8 contains P63

(B9-2206).

Focusing on Peaks#3-#6, in the Cl-8 homolog group, their vapour pressures are similar (ranging from

0.0017-0.0019 Pa [34] so differences in proportions of the peaks are probably not due to volatility. Relative areas

of these four peaks are compared in Figure 3, along with a comparison to air from the southern USA, a past usage

region of toxaphene. Based on relative peak areas, normalized to Peak#4, Peaks#3 and #4 were significantly (p

<0.05) depleted in air and water compared to the standard, and Peak#3 was depleted in the water compared to the

air (p<0.05), but no significant differences in the ratio of these congeners were seen in the air samples between

August and May. Shoeib et al. [24,33] found similar depletions in air samples from Point Petre, Lake Ontario.

P39 (Peak#3) and P42 (Peak#4) are especially labile in the environment [26]. Peak#3 and Peak#4 were also

depleted in soil and air samples from Alabama [19] and in sewage sludge [35]. Also based on relative peak areas,

Peak#6 is slightly enriched in the air and significantly depleted in the water (p<0.05). Air sampled in the southern

USA also showed an enrichment of Peak#6. Even though Alabama air and Lake Superior air and water show similar

depletions of Peak#3 and Peak#4, the pattern is quite different, see Figure 3. The pattern in Lake Superior water is

reflected in the overlying air, which may imply air-water gas exchange has a larger contribution to the overlying air

than long range transport from the southern USA or current usage. This is reinforced by a net flux direction that

predicts volatilization from the surface of Lake Superior to the overlying air (see next section). Peak#8 was found in

the water but not in the air, and based on relative peak areas, shows depletions in the water compared to the standard.

126

Table 3: Fugacity Ratio and Flux Calculations for Cruises and Annual Predictions, see Figure 1 for sample locations.

Month Average Temperature Air Average H Fugacities Wind Mass Transfer Coefficients Fraction Flux

Water f w f a f w/f a U 10 k a k w k OL Open (ng/m2d)

T a (K) T w (K) C a (ng/m3) C w ng/L Pa m3/mol Pa Pa m/s m/s m/s m/s Water N v N d N

Cruises

August 1996

SSM - station (st) 80 287 284 0.028 0.92 0.13 2.9E-10 1.6E-10 1.8±0.6 5.2 4.8E-03 9.0E-06 4.7E-03 1.0 23 12 11±6.5

st 80 - st 139 282 280 0.017 0.92 0.09 2.1E-10 9.6E-11 2.2±0.7 5.6 5.1E-03 1.0E-05 5.0E-03 1.0 18 7.4 10±5.9

st 139 - st 169 288 286 0.024 0.92 0.17 3.7E-10 1.4E-10 2.6±0.9 4.1 3.9E-03 5.9E-06 3.9E-03 1.0 24 8.1 16±9.3

st 169 - st 51 290 285 0.041 0.92 0.15 3.3E-10 2.4E-10 1.4±0.4 4.4 4.2E-03 6.8E-06 4.2E-03 1.0 23 15 7.9±4.6

Mean 287 284 0.028 0.92 0.13 3.0E-10 1.6E-10 2.0±0.6 4.8 4.5E-03 7.9E-06 4.5E-03 1.0 22 11 11±6.6

May 1997

SSM - Thunder Bay 275 275 0.0049 0.92 6.0E-02 1.3E-10 2.7E-11 4.9±1.5 4.4 4.2E-03 6.7E-06 4.1E-03 1.0 8.6 1.8 6.9±3.9

Thunder Bay - Duluth 276 275 0.011 0.92 6.0E-02 1.3E-10 6.1E-11 2.2±0.6 3.6 3.6E-03 4.8E-06 3.6E-03 1.0 7.4 3.4 4.0±2.3

Duluth - st 157 277 276 0.012 0.92 6.7E-02 1.5E-10 6.7E-112.2±0.6 4.2 4.0E-03 6.2E-06 4.0E-03 1.0 9.2 4.1 5.1±2.9

st 157 - Thunder Bay 277 275 0.017 0.92 6.0E-02 1.3E-10 9.4E-11 1.4±0.4 5.0 4.6E-03 8.2E-06 4.6E-03 1.0 9.5 6.7 2.8±1.6

Thunder Bay - SSM 278 275 0.015 0.92 6.0E-02 1.3E-10 8.3E-11 1.6±0.5 2.3 2.7E-03 2.3E-06 2.7E-03 1.0 5.6 3.5 2.1±1.2

Mean 277 278 0.012 0.92 6.2E-02 1.4E-10 6.7E-11 2.5±0.7 3.9 3.8E-03 5.7E-06 3.8E-03 1.0 8.1 3.9 4.2±2.4

Annual Predictions Monthly Temperature Monthly Air

C a (ng/m3)

January 265 275 0.0035 0.92 5.4E-02 1.2E-10 1.9E-11 6.4 7.5 7.2E-03 1.9E-05 7.1E-03 0.9 14 2.1 10

February 264 274 0.0031 0.92 5.2E-02 1.2E-10 1.7E-11 7.0 6.86.5E-03 1.7E-05 6.4E-03 0.44 12 1.7 4.6

March 269 274 0.0049 0.92 4.9E-02 1.1E-10 2.6E-11 4.1 6.0 5.7E-03 1.4E-05 5.7E-03 0.4 10 2.4 3.0

April 275 273 0.0088 0.92 4.8E-02 1.1E-10 4.9E-11 2.2 5.6 5.3E-03 1.3E-05 5.3E-03 0.9 8.9 4.1 4.4

May 280 274 0.015 0.92 5.1E-02 1.1E-10 8.1E-11 1.4 4.9 4.7E-03 1.2E-05 4.7E-03 1.0 8.2 5.8 2.3

June 286 277 0.026 0.92 6.7E-02 1.5E-10 1.5E-10 1.0 4.6 4.4E-03 1.1E-05 4.4E-03 1.0 9.8 9.7 0.1

July 287 279 0.029 0.92 8.3E-02 1.8E-10 1.7E-10 1.1 5.0 4.8E-03 1.2E-05 4.8E-03 1.0 13 12 1.4

August 292 285 0.043 0.92 1.4E-01 3.1E-10 2.5E-10 1.2 4.6 4.4E-03 1.1E-05 4.3E-03 1.0 20 16 3.7

September 289 286 0.032 0.92 1.6E-01 3.6E-10 1.8E-10 1.9 5.25.0E-03 1.2E-05 4.9E-03 1.0 26 13 13

October 282 281 0.018 0.92 1.0E-01 2.2E-10 9.9E-11 2.2 7.3 7.0E-03 1.8E-05 6.9E-03 1.0 23 10 13

November 271 279 0.0064 0.92 7.9E-02 1.8E-10 3.5E-11 5.1 7.57.1E-03 1.8E-05 7.0E-03 1.0 20 3.9 16

December 268 278 0.0044 0.92 7.4E-02 1.6E-10 2.4E-11 7.0 7.67.3E-03 1.9E-05 7.2E-03 1.0 19 2.7 16

127

Figure 2:Chromatograms of total toxaphene and Cl-7 to Cl-9 homolog groups, top = air, middle = water and bottom = standard. Peak#1 = P26, #2 = B8-1412 [22] (no Parlar number), #3 = P39, #4 = P40+P41, #5 = P42, #6 = P44, #7 = P50 and #8 = P63. P26 appears lower than actually present in air and water samples because it splits between silicic acid fractions 1 and 2. Fraction 2 is shown here.

Figure 3: Averaged relative proportions of Parlar congeners Peak#3, Peak#5 and Peak#6 normalized to Peak#4 (=1.00) for air and water samples.

STD

Air

Water

Cl-7 Homolog

Total ToxapheneCl-8 Homolog

Cl-9 Homolog

STD

Air

Water

STD

Air

Water

STD

Air

Water

4 65

3

4

65

3

46

5

3

1

2

12

1

2

7

8

7

8

7

8

Standard

LS 96 AIR

LS 97 AIR

LS WATER

Southern USA0.0

0.2

0.4

0.6

0.8

1.0

1.2

Peak#3 Peak#5 Peak#6

Nor

mal

ized

to P

eak#

4 A

rea

128

Long range transport of toxaphene to the Great Lakes is still likely to occur, Li et al. [36] estimated that

29kt of toxaphene residues in the U.S.A. agricultural soils of which 360t were emitted to the air in 2000. Harner et

al., [37] modelled soil to air exchange of toxaphene in Alabama and estimated that volatilization from agricultural

soils could supply 3000-11 000 kg/y of toxaphene to the atmosphere.

Temperature Dependence

The relationship of atmospheric toxaphene to air temperature was investigated by plotting Log P (partial

pressure, Pa) versus 1/Ta. The slope and intercept were : m = - 3291 and b = 1.67, r2 = 0.76. Slopes found for

similar plots in other recent studies are: South Carolina -2583 [30], Eagle Harbor -2438 [28], Point Petre -2284

[24] and Lakes Superior and Michigan -3989 [10]. Although the toxaphene concentrations in these locations

differed by up to 50-fold, the slopes are quite similar. However in the 1996-1997 Alabama study, no significant

temperature dependence was found [19] and it was assumed that other factors were controlling atmospheric

concentrations, e.g. soil moisture since volatilization may be suppressed when the soil is dry.

Gas Exchange Estimates from Parallel Air and Water Samples

Gas exchange dominates loadings of OCPs to the Great Lakes [1,2]. Unlike wet and dry deposition, gas

exchange is a reversible process where deposition and volatilization occur simultaneously. Gas exchange of

toxaphene in August 1996 and May 1997 was estimated using the modified two-film model with fugacity

definitions [38]:

fw = 10-9 Cw H/M (1)

fa = 10-9 Ca RTa/M (2)

fw/fa = CwH/CaRTa (3)

In equations 1-3, fw and fa are the fugacities in water and air, Cw and Ca are the dissolved concentration in water and

vapour-phase concentration in air (both ng/m3), H is the Henry's Law constant at the temperature of the water (Pa

m3/mol), R is the gas constant ( 8.314 Pa m3/mol K), Ta is the temperature of the air (K), and M is the molecular

weight (414 for octachlorobornane). Temperatures used here are those measured on board the ship at a height of 10 m

for air and a depth of 2 m for water. The Henry's Law constant for technical toxaphene was calculated using the

measured water temperatures from equation 4 [39].

Log H = 10.42 - 3209/Tw (4)

Values of fw/fa < 1 and > 1 imply net deposition and volatilization respectively; fw/fa = 1 is thermodynamic

equilibrium. The fugacity ratios ranged from 1.4-2.6 in August and 1.3-4.7 in May. The error associated with the

fugacity ratios for discrete air events was ±30%, propagated from the relative standard deviations of the water

concentration (±22%) and the Henry’s Law constant (±20%). Mean fugacity ratios were 2.0 ± 0.6 in August and

2.3 ± 0.7 in May and were significantly different from 1 (p<0.05).

Estimate of Fluxes

Fluxes (N) were estimated using the following equations:

Deposition ND (ng /m2 d)= 109 M Daw fa (5)

Volatilization NV (ng /m2 d)= 109 M Daw fw (6)

129

Daw (mol/ m2 d Pa) = 86400KOG/RTa (7)

waaOG k RT

H +

k

1 =

1

K (8)

Where Daw is the transport capacity, KOG is the overall mass transfer coefficient expressed on a gas-side basis (cm/s)

and ka and kw are individual mass transfer coefficients for air and water. The mass transfer coefficients were calculated

using equations 9 and 10 from Galarneau et al. [40] which are simplified from Hornbuckle et al. [41]:

k (cm / s) = 15(0.2 U + 0.3)(1/ M + 1/ 29)

( V + 19.7a 10

0.5

a1/3Σ 1 3 2

0 61

/

.

)

(9)

k (cm / h) = 0.45 U V

29.6w 10

1.64 m

−0 3.

(10)

In equations 9 and 10, ΣVa is the sum of atomic diffusion volumes for toxaphene, calculated from the incremental

volumes in Table 11-1 of ref [42], Vm is the molar volume (cm3/mol) and U10 is the wind speed at 10 m (m/s). Mass

transfer coefficients were also calculated from relationships in Mackay and Yuen [43]. These values of kw were 40-

60% larger and 2-5% smaller for ka. Since the Henry’s Law constant for toxaphene is relatively low compared to

other OCPs, over 95% of the resistance to transfer lies in the air film and and KOG is dominated by the ka. Thus the

KOG values calculated by the two approaches differ by <10%.

Table 3 presents deposition, volatilization and net fluxes of toxaphene to Lake Superior for the cruise in the

months of August and May and predicted fluxes for the other months (see below). Net fluxes were from water to air;

these averaged 8.6 ± 5.0 ng/m2 d in August 1996 and 3.7 ± 2.2 ng/m2 d in May 1997, where the net flux >1 is

volatilization and <1 is deposition. The error in the water and air concentrations, Henry’s Law constant and the mass

transfer coefficient were propagated yielding an accumulated error of 58%, the uncertainty in the mass transfer co-

efficient was assumed to be 30% [43].

James et al. [10] calculated higher net fluxes for the spring (25-33 ng/m2 d) and summer (8.3-19 ng/m2 d),

Swackhamer also calculated a higher annually averaged flux of 13 ng/m2 d the major difference between these three

sets of calculations is the mass transfer coefficient which is a function of windspeed. Swackhamer et al. [3] and James

et al. [10] assumed much higher wind speeds of 7-11 m/s after Hoff et al [44], where the windspeeds used here were

hourly measurements taken on board the ship and averaged over the time of the air sample, averaging 4.8 m/s in May

and 3.9 m/s in July.

Estimation of Annual Gas Exchange Loadings

Monthly averaged atmospheric concentrations of toxaphene were estimated from the Log P versus 1/T

relationship (see above). For 1996 and 1997, daily water temperatures were obtained from National Oceanic and

Atmospheric Administration (NOAA), Great Lakes Environmental Research Laboratory and hourly buoy (over the

lake) air temperature and wind speed data were obtained from NOAA National Weather Service. These data were

averaged by month and used to estimate monthly air concentrations and toxaphene fluxes. Predicted monthly

averaged atmospheric concentrations ranged from 3.1 - 43 pg/m3 (Table 3). The water concentration of toxaphene

130

was assumed constant over the year, since concentrations were not statistically different between August, 1996 and

May, 1997.

Results of the monthly fugacity ratio and flux calculations are given in Table 3 and Figure 4 for a year

representative of 1996 and 1997; also shown for comparison are the fugacity ratios calculated from the parallel air

and water data on the two cruises. The fugacity ratios calculated from the cruise data are higher than those

predicted from the annual air cycle, August cruise = 2.0 versus predicted = 1.2 and May cruise = 2.3 versus

predicted = 1.4. Lower measured air concentrations are due to lower air temperatures during the cruises than the

monthly averages. The monthly fugacity ratios ranged from 1.0-7.0 and indicate the potential of toxaphene to

volatilize from the lake. The fugacity ratios were highest in the winter (despite colder water temperatures) and

lowest in the summer, driven by atmospheric concentrations that are approximately 4 times lower in the winter than

summer.

Figure 4: Monthly fugacity ratios (bars), air and water temperatures (solid and dashed lines) for Lake Superior.

Figure 5: Monthly toxaphene fluxes (bars) and windspeed (solid line) for Lake Superior.

The flux was also calculated on a monthly basis for the representative year. Inhibition of gas exchange due to

ice cover was included since Lake Superior is partially frozen in winter. Lake Superior is 60-65% ice covered

during February and March, 10% in January and April and ice free for the rest of the year, (Table 3). Deposition,

Janu

ary

Febru

ary

March

April

May

June Ju

ly

Augus

t

Septem

ber

Octobe

r

Novem

ber

Decem

ber

0

1

2

3

4

5

6

7

260

265

270

275

280

285

290

295

Predicted Paired Tw Ta

Temperature (K

)

f /f

wa

Janu

ary

Febr

uary

Mar

ch

April

May

June

July

Augu

stSe

ptem

ber

Oct

ober

Nov

embe

rD

ecem

ber

0

5

10

15

20

0

2

4

6

8

Predicted Paired Windspeed

Net

Flu

x (n

g/m

d)

2

Windspeed (m

/s)

131

volatilization and net fluxes are given in Table 3 and net values are shown in Figure 5. Net fluxes ranged from 0.1-

16 ng/m2d, with higher fluxes from September to January (10-16 ng/m2d) than February to August (0.1-4.6

ng/m2d).

The parallel air and water samples in this study provide the most accurate data for calculating fugacity

ratios because the analysis of both air and water was done by the same method and in the same laboratory.

Fugacity ratios were also estimated from airborne toxaphene data at Eagle Harbor for 1996-1997 [28] and water

measurements from this study. The pattern in the fugacity ratios was the same as for the data presented above, but

since the Glassmeyer et al. [28] atmospheric concentrations are lower (annual average of 4.8 ± 4.4 pg/m3), the

fugacity ratios were larger (varying from 7-18). Swackhamer et al. [3] predict a similar seasonal cycling using

predicted air concentrations based on 1988-89 air measurements [15]. They also estimated higher rates of

volatilization in the winter than spring or fall and net absorption of toxaphene during the summertime. James et al. [10]

showed a similar estimate of toxaphene in Lakes Michigan and Superior, with percent saturation of 200-1000%

(fugacity ratio of 2-10).

The area and volume of Lake Superior are 82 100 km2 and 12 100 km3. Assuming uniform toxaphene

concentrations with depth, the lake is estimated to contain approximately 11 000 kg of toxaphene. We predict a net

volatilization loss of approximately 220 kg toxaphene over the year from Lake Superior, or approximately 2% of the

total toxaphene in the lake. Swackhamer et al. [3] and James et al. [10] predicted a higher net volatilization of 389-

650 kg/year or 3.6-6.0 % of the lake burden per year. As discussed earlier, the major difference between three sets of

flux calculations is the mass transfer coefficient used, which is a function of wind speed. Swackhamer et al. [3] and

James et al. [10] assumed higher wind speeds of 7-11 m/s, after Hoff et al. [44], where for the annual loadings, we

used monthly averaged windspeeds from a buoy over the lake. From these three studies we can conclude that

toxaphene is volatilizing from Lake Superior and will continue to do so for many decades. As emissions and the

global inventory of toxaphene decline, sources such as Lake Superior will only increase their net flux to the

atmosphere.

Acknowledgement

We thank Janine Wideman and the crew and technical operations of the Canadian Coast Guard Ship Limnos for

help during sampling and Walter Vetter, University of Jena, Germany, for supplying a B8-1413 standard.

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4. Hainzl D, Burhenne J, Parlar H. 1994. Theoretical consideration of the structure variety in the toxaphene mixture taking into account recent experimental results. Chemosphere 28:245-251. 5. Li YF. 2001. Toxaphene in the United States, 1. Usage gridding. J Geophys Res 106:17919-17927. Li YF, Bidleman TF, Barrie LA. 2001. Toxaphene in the United States, 2. Emissions and residues. J Geophys Res 106:17929-17938. 6. Rapaport RA, Eisenreich SJ. 1986. Atmospheric deposition of toxaphene to eastern North America derived from peat accumulation. Atmos Environ 20:2367-2379. 7. Donald DB, Stern GA, Muir DCG, Fowler BR, Miskimmin BM, Bailey R. 1998. Chlorobornanes in water, sediment and fish from toxaphene treated and untreated lakes in western Canada. Environ Sci Technol 32:1391-1397. 8. Howdeshall MJ, Hites RA.1996. Historical input and degradation of toxaphene in Lake Ontario sediment. Environ Sci Technol 30:220-224. 9. Pearson R, Swackhamer D, Eisenreich S, Long D. 1997. Concentrations, accumulation and inventories of toxaphene in sediments of the Great Lakes. Environ Sci Technol 31:3523-3529. 10. James RR, McDonald JG, Symonik DM, Swackhamer DL, Hites RA. 2001. Volatilization of toxaphene from Lakes Michigan and Superior. Environ Sci Technol 35: 3653-3660. 11. Glassmeyer ST, De Vault DS, Meyer TR, Hites RS. 1997. Toxaphene in Great Lakes fish: A temporal, spatial and trophic study. Environ Sci Technol 31:84-88. 12. Glassmeyer ST, De Vault DS, Meyer TR, Hites RS. 2000. Rates at which toxaphene concentrations decrease in lake trout from the Great Lakes. Environ Sci Technol 34:1851-1855. 13. Ontario Ministry of the Environment, 1999.Guide to Eating Ontario Sport Fish: 1999-2000, 20th ed., Canada, Queens printer for Ontario. 14. Cortes DR, Basu I, Sweet CW, Brice KA, Hoff RW, Hites RA. 1998. Temporal trends in gas phase concentrations of chlorinated pesticides measured at the shores of the Great Lakes. Environ Sci Technol 32:1920-1927. 15. Hoff RM, Muir DGC, Grift NP. 1992. The annual cycle of polychlorinated biphenyls and organochlorine pesticides in air in southern Ontario: 1 Air concentration data. Environ Sci Technol 26:266-275. 16. Mackay D, Bentzen E. 1997. The Role of the Atmosphere in Great Lakes Contamination. Atmos Environ 31: 4045-4047. 17. Jantunen L, Bidleman T. 1998. Organochlorine pesticides and enantiomers of chiral pesticides in Arctic Ocean water. Arch Environ Toxicol Contam 35:218-228. 18. Bidleman TF, Falconer RL, Walla M. 1995. Toxaphene and other organochlorine compounds in air and water at Resolute Bay, N.W.T., Canada. Sci Total Environ 160/161:55-63. 19. Jantunen L, Harner T, Bidleman TF. 2000. Toxaphene, chlordane and other organochlorine pesticides in Alabama air. Environ Sci Technol 34:5097-5105. 20. Frenzen G, Hainzl D, Burhenne J, Parlar H. 1994. Structure elucidation of the three most important toxaphene congeners by X-ray analysis. Chemosphere 28:2067-2074.

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21. Stern GA, Muir DCG, Ford CA, Grift NP, Dewailly E, Bidleman TF, Walla MD. 1992. Isolation and identification of two major recalcitrant toxaphene congeners in aquatic biota. Environ Sci Technol 26:1838-1840. 22. Andrew P, Vetter W. 1995. A systematic nomenclature system for toxaphene congeners Part 1: Chlorinated bornanes. Chemosphere 31:3879-3886. 23. Saleh MA. 1991. Toxaphene chemistry, biology, toxicity and environmental fate. In Ware. GW, ed, Reviews of Environmental Contamination and Toxicology., Springer-Verlag, New York, USA. 118, p1-85. 24. Shoeib M, Brice KA, Hoff R. 1999. Airborne concentrations of toxaphene congeners at Point Petre (Ontario) using gas chromatography-electron capture negative ion mass spectrometry (GC-ECNIMS). Chemosphere 5:849-871. 25. Nikiforov VA, Karavan VA, Mitlsov SA. 1999. Composition of Camphechlor. Organohalogen Compounds 41:601-604. 26. Vetter W, Scherer G. 1999. Persistency of toxaphene components in mammals that can be explained by molecular modelling. Environ Sci Technol 33:3458-3461. 27. McConnell LL, Bidleman TF, Cotham WE, Walla MD. 1998. Air concentrations of organochlorine insecticides and polychlorinated biphenyls over Green Bay, Wisconsin and the four lower Great Lakes. Environ Pollut 101:391-399. 28. Glassmeyer ST, Brice KA, Hites RA. 1999. Atmospheric concentrations of toxaphene on the coast of Lake Superior. J Gt Lakes Res 25:492-499. 29 Cussion S, Bidleman TF, Jantunen LMM, 2001. Interlaboratory Study of Toxaphene in Ambient Air, Ontario Ministry of the Environment report to the Meteorological Service of Canada, Downsview, ON, Canada. 30. Bidleman TF, Alegria H, Ngabe B, Green C. 1998. Trends of chlordane and toxaphene in ambient air of Columbia, South Carolina. Atmos Environ 32:1849-1856. 31. Alegria HA, Bidleman TF, Shaw TJ. 2000. Organochlorine pesticides in the ambient air of Belize,Central America. Environ Sci Technol 34:1953-1958. 32. Jantunen LM. 1997. Air-Water Gas Exchange of Toxaphene in Arctic Regions, Masters Thesis, University of Toronto, Toronto, Ontario, Canada. 33. Shoeib M, Brice KA, Hoff RM. 2000. Studies of toxaphene in technical standard and extracts of background air samples (Point Petre, Ontario) using multidimensional gas chromatography-electron capture detection (MDGC-ECD). Chemosphere 40:201-211. 34. Bidleman TF, Leone AD, Falconer RL. 2002. Vapour pressures and heats of vapourization of toxaphene congeners. J Chem Eng Data submitted. 35. Buser HR, Müller MD. 1995. Isomer and enantioselective degradation of hexachloro-cyclohexane isomers in sewage sludge under anaerobic conditions. Environ Sci Technol 29:664-672. 36. Li YF, Bidleman TF, Barrie LA. 2001. Toxaphene in the United States, 2. Emissions and residues. J Geophys Res 106:17929-17938. 37. Harner T, Bidleman TF, Mackay D. 2001. Soil-air exchange model of persistent pesticides in the U.S. Cotton Belt. Environ Toxicol Chem 20: 1612-1621. 38. Mackay D. 1991. Multimedia Environmental Models: Fugacity Approach. Lewis, Chelsea, MI, USA.

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39. Jantunen LM, Bidleman TF. 2000. Temperature Dependent Henry’s Law Constant for Technical Toxaphene, Chemosphere- Global Change Science 2:225-231. 40. Galarneau E, Audette CV, Bandemehr A, Basu I, Bidleman TF, Brice KA, Burniston DA, Chan CH, Froude F, Hites RA, Hulting ML, Neilson M, Orr D, Simcik MF, Strachan WMJ, Hoff R. 2000. Atmospheric deposition of toxic substances to the Great Lakes: IADN results to 1996. Environment Canada and US EPA, 2000, ISBN 0-662-29007-0, Downsview, ON, Canada. 41. Hornbuckle KC, Jeremiason JD, Sweet CW, Eisenreich SJ. 1994. Seasonal variations in air-water exchange of polychlorinated biphenyls in Lake Superior. Environ Sci Technol 28:1491-1501. 42. Reid RC, Prausnitz JM, Poling BE. 1987. The Properties of Gases and Liquids, 4th ed., McGraw-Hill, New York, NY, USA, 741 pp. 43. Mackay D, Yuen ATK. 1983. Mass transfer coefficient correlations for volatilization of organic solutes from water. Environ Sci Technol 17:211-217. 44. Hoff RM, Bidleman TF, Eisenreich SJ. 1993. Estimates of PCC Loadings from the atmosphere to the Great Lakes. Chemosphere 27:2047-2055.

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Paper 5

Air-Water Gas Exchange of Hexachlorocyclohexanes (HCHs)

and the Enantiomers of αααα-HCH in Arctic Regions


Journal of Geophysical Research, 1996. 101, 28837-28846.

The version here includes corrections published 1997, ibid. 102, 19279-19282.

Contributions: Liisa Jantunen and Terry Bidleman jointly collected, processed and analysed samples on the

BERPAC'93 expedition. In 1994, Liisa collected, processed and analysed samples from the AOS’94 cruise. The paper

was written jointly by Liisa and Terry.

136

Abstract

In the summers of 1993 and 1994 air and water samples were taken in the Bering-Chukchi seas and on a transect

across the polar cap to the Greenland Sea, to measure the air-sea gas exchange of hexachlorocyclohexanes (HCH) and

the enantiomers of α-HCH. Atmospheric concentrations of α- and γ-HCH have decreased by three-fold or more since

the mid-1980s, whereas concentrations in surface water have shown little change. The saturation state of surface water

(water/air fugacity ratio) was determined from the air and water concentrations of HCHs and their Henry's law constants

as a function of temperature. Fugacity ratios >1.0 indicated net volatilization of α-HCH in all regions except the

Greenland Sea, where concentrations in air and water were close to equilibrium. Net deposition of γ-HCH in the

Chukchi Sea was indicated by fugacity ratios <1.0. In other regions γ-HCH was volatilizing or near air-water

equilibrium. Enantioselective degradation of (-)α-HCH was found in surface water of the Bering-Chukchi seas. The ER

was reversed in the Canada Basin and Greenland Sea, where (+)α-HCH was preferentially lost. The same order of

enantioselective degradation was seen in air within the marine boundary layer of these regions, which provides direct

evidence for sea-to-air transfer of α-HCH.

Introduction

Atmospheric transport has led to extensive contamination of the Arctic by organochlorine compounds (OCs),

including polychlorinated biphenyls (PCBs) and the pesticides toxaphene, chlordane, DDT, and

hexachlorocyclohexanes (HCH) (Barrie et al., 1992; Bidleman et al. 1995a; Fellin et al., 1996; Oehme et al., 1995;

Patton et al., 1989, 1991). The latter are the most abundant OCs in air and water of arctic and subarctic regions

(Bidleman et al., 1995a; Fellin et al., 1996; Hargrave et al., 1988; Iwata et al., 1993a). Technical HCH is an insecticidal

mixture of 60-70% α-HCH, 5-12% β-HCH, 10-15% γ-HCH (Iwata et al., 1993a), and minor percentages of other

isomers, which was introduced during World War II. Canada, the United States and most European countries have

eliminated the use of technical HCH in favor of lindane (pure γ-HCH), but technical HCH was heavily used in Asian

countries throughout the 1980s (Barrie et al., 1992; Hinckley et al., 1991; Voldner and Li, 1995).

In a process known as "global distillation" (Goldberg, 1975), HCHs volatilize quickly from temperate and tropical

regions (Takeoka et al., 1991) and are atmospherically transported worldwide. HCHs accumulate in water bodies,

largely through air-water gas exchange. A global model of HCH circulation estimates that over 90% of the HCH

present in the environment at the end of 1985 was held in the world's oceans (Strand and Høv, 1994). Vapor pressure

and Henry's law constants of HCHs decrease at lower temperatures, favoring partitioning of HCHs into condensed

phases in colder climates. Even though the heaviest usage of HCHs has been in tropical and subtropical regions, levels

in surface seawater are an order of magnitude higher in northern oceans (Hargrave et al., 1988; Iwata et al., 1993;

Schreitmüller and Ballschmiter, 1995). Concentrations of HCHs in tree bark increases with latitude, being relatively

high in Canada, Nordic countries, Alaska and Russia and lower in tropical countries (Simonich and Hites, 1995). Thus,

the transport and global distribution of HCHs exemplifies the "cold condensation effect" (Mackay and Wania, 1995;

137

Wania and Mackay, 1993, 1995). The Arctic was once thought to be a permanent sink for OCs but recent

measurements in subarctic and arctic regions have shown that HCHs are now revolatilizing in response to decreasing

atmospheric concentrations. (Bidleman et al., 1995b; Falconer et al., 1995a; Jantunen and Bidleman, 1995).

The pesticide α-HCH is a chiral compound that is present as a racemate in the technical HCH mixture. Abiotic

processes that degrade α-HCH, for example hydrolysis and photolysis, are not chirally selective, but biological systems

have the ability to discriminate between the two enantiomers. Microbial attack (Buser and Müller, 1995; Faller et al.,

1991a; Hühnerfuss, 1992; Ludwig et al., 1992; Möller and Hühnerfuss, 1993), enzymatic pathway in higher organisms

(Hummert et al., 1995; Möller and Hühnerfuss, 1993; Mössner et al., 1992; Müller et al., 1992; Tanabe et al., 1996) and

differential transport across biological membranes (Möller and Hühnerfuss, 1993) can alter the relative proportions of

the enantiomers. Enantioselective breakdown of α-HCH has been reported in water from the Canadian Arctic (Falconer

et al., 1995a,b) and the North Sea (Faller et al., 1991a,b; Möller and Hühnerfuss, 1993; Pfaffenberger et al., 1992), soils

from Germany (Falconer et al., 1995c; Müller et al., 1992), anaerobic sewage sludge (Buser and Müller, 1995) and

birds, mammals and fish (Möller and Hühnerfuss, 1993; Mössner et al., 1992; Müller et al., 1992; Pfaffenberger et al.,

1994; Tanabe, et al., 1996).

In August to September 1993, water and air samples were collected in the Bering and Chukchi seas on a U.S.-

Russian cruise (BERPAC-93) (Jantunen and Bidleman, 1995). Samples were again collected on a Canadian-U.S.

transect of the Arctic Ocean in July to September 1994 (Arctic Ocean Sections, AOS-94). The purpose of these

measurements was to investigate the enantioselective degradation of α-HCH and air-sea exchange of α- and γ-HCH.

Results of the BERPAC-93 gas exchange studies have been detailed elsewhere (Jantunen and Bidleman, 1995). Here

we report the air-water gas exchange from the high Arctic, the enantiomeric composition of α-HCH in air and water

from the arctic and subarctic regions and the use of enantiomers to follow gas exchange.

Experimental Section


In the summer of 1993, the Russian ship R/V OKEAH completed a 50 day cruise (BERPAC-93) of the Bering and

Chukchi seas where water and air sampling for HCHs was done. The cruise track is shown in a previous publication

(Jantunen and Bidleman, 1995). An extensive survey was done from the Aleutian Islands to the ice edge at 74oN, on

both the American and Russian sides of the Bering and Chukchi seas. Air and water temperatures ranged from 11.0 to

0.7oC. Most sampling was done on the continental shelf with bottom depths of less than 75 m, except for a few stations

in the Aleutian Basin of the Bering Sea and the north Chukchi Sea at the ice edge where maximum bottom depths were

~2000 m.

In July to September 1994 another series of samples was collected from the Canadian research vessel CCGS Louis

S. St. Laurent on a transect of the Arctic Ocean and adjacent waters from Victoria, British Columbia to Halifax, Nova

Scotia across the North Pole. The cruise track is shown in Figure 1. The areas where water samples were collected

138

included the Chukchi Sea (stations 1-7), the Canada Basin (stations 8-30), the Eurasia Basin (station 31-38) and the

Greenland Sea (station 39). Bottom depths were mainly >1000 m. Water temperatures ranged from 4 to -1.7oC, but

most surface water temperatures at high latitudes were near -1.7oC, the freezing point of sea water. Water samples were

taken at varying depths by using 12 L Teflon-lined Go-Flo bottles (General Oceanics) positioned on a rosette that was

equipped with a conductive and temperature at depth (CTD) probe to collect the hydrographic information. Surface

samples were collected by a submersible pump; the line running from the pump was Teflon tubing surrounded by a

flexible metal mesh. Samples were transferred to pressurizable stainless steel cans using Teflon tubing, then passed

through a glass fiber filter followed by a 1-g C8-bonded silica solid phase extraction cartridge (SPE) to isolate HCHs.

Air samples were collected continuously when the ship was in motion, using a glass fiber filter-polyurethane foam (PUF)

trap (Jantunen and Bidleman, 1995).

Water and air sampling media were processed for analysis by previously described methods (Jantunen and

Bidleman, 1995). Extract volumes were brought to 1 mL (quantitative analysis) or 100 µL (enantiomer analysis) by

blowdown with nitrogen into isooctane.

Figure 1: Sampling and cruise track, from BERPAC-93 (solid line) and Arctic Ocean Section’94 (dashed), with extent of ice cover for August 1994 (dotted line). Small number indicate locations of some sampling stations.

Eurasia Basin

Canada Basin

Chukchi Sea38

35

27

2518 39

Greenland SeaBering Sea

Eurasia Basin

Canada Basin

Chukchi Sea38

35

27

2518 39

Greenland SeaBering Sea

139

Analysis

Negligible amounts of HCHs were found on water and air filters (Jantunen and Bidleman, 1995), so quantification

and enantiomeric analysis were carried out on extracts of SPE cartridges and PUF plugs. Instrumentation and

procedures for quantitative analysis were detailed earlier (Jantunen and Bidleman, 1995).

Enantiomeric analysis of α-HCH was done by capillary GC-electron capture negative ion mass spectrometry using

a Hewlett-Packard 5890 GC-5989B MS Engine. The primary column used to separate the two enantiomers was fused

silica coated with 30% tert-butyldimethylsilylated betacyclodextrin in PS-086, referred to as BSCD (BGB Analytik AG,

Switzerland, 20 m length, 0.25 mm i.d., 0.25 µm film thickness). The temperature program was: initial temperature

90oC, 15oC min-1 to 140oC, 1.0oC min-1 to 180oC; hold for 2.0 min, 20oC min-1 to 240oC; hold for 5.0 min. The

confirmational column was fused silica containing 20% permethylated betacyclodextrin in polydimethylsiloxane,

referred to as Betadex (Supelco, 30 m length, 0.25 mm i.d., 0.25 µm film thickness). The temperature program for the

Betadex column was: initial temperature 90oC, 15oC min-1 to 130 oC, 1.0oC min-1 to 190oC, 20oC min-1 to 230oC; hold

for 5.0 min. The two columns employed here reverse the elution order of the enantiomers. On Betadex (+)α-HCH elutes

first, as determined with a standard that had been depleted in (+)α-HCH by reaction with brucine (Falconer et al.,

1995b). Comparison of the two columns for non racemic samples shows that (-)α-HCH elutes first on BSCD (this work

and Falconer et al., 1995c). The enantiomeric ratio (ER) of α-HCH is defined as the area of (+)α-HCH/(-)α-HCH.

Samples were injected splitless (2µL, split opened after 0.5 minutes). Other instrument conditions were: helium carrier

gas at 50 cm s-1 for BSCD and 40 cm s-1 for Betadex, injector temperature 250oC, transfer line temperature 250oC, ion

source 150 oC and quadrupole 100oC. The mass spectrometer was operated in the negative ion mode with methane at

~1 torr. The ions monitored were 255 and 257, and also 253 for some samples. Ion 255 is the most abundant and was

used in calculating ER results, 253 and 255 were used as qualifiers.


Quality Control

A summary of blank and recovery experiments for the AOS-94 cruise is presented here. Similar results were

obtained on BERPAC-93 (Jantunen and Bidleman, 1995). Water blanks were done while on the ship by passing 100 mL

of deionized water through a filter and SPE cartridge. Average blank values for the cartridges were 0.18 ± 0.03 ng (±

indicates standard deviation) for α-HCH and 0.06 ± 0.01 ng for γ-HCH, n=4. Recovery experiments were done by

spiking surface seawater with 15 ng L-1 of α-HCH and 5 ng L-1 of γ-HCH. Percent recoveries were 71 ± 9 % for α-HCH

and 74 ± 13 % for γ-HCH, n=5, after correcting for the blank value and native amount in seawater. Reported sample

concentrations were corrected for average recoveries. The amount of α-HCH and γ-HCH on sample and spike filters

was <0.05 ng (the detection limit). Duplicate 4-L samples were processed at four stations, the average differences were

4 % for α-HCH and 10 % for γ-HCH. Field blanks were done for air on two occasions by placing a clean PUF and filter

into the air sampling apparatus and drawing air for 30 seconds. Blanks for the PUF were 0.7-0.8 ng α-HCH and 0.25-

140

0.55 ng γ-HCH. The blanks for the filters were <0.05 ng for both α-HCH and γ-HCH. Recoveries of 11-23 ng, α-HCH

and γ-HCH from two spiked PUF plugs were 80-89 % and 96-97 % , respectively.

The degree of breakthrough was determined by analyzing the front and back PUFs separately. When averaged

according to air temperature, breakthrough results (as percent of front PUF value) were 16 ± 14 % for α-HCH and 5 ± 8

% for γ-HCH at >0oC, n=9 and decreased to 3 ± 2 % for α-HCH and 3 ± 3 % for γ-HCH at <0oC, n=5. The sum of the

back and front PUF quantities was used to calculate the concentration of HCHs in air, but only front PUFs were used for

enantiomeric analysis.

Confirmational analysis for enantiomers was done by comparing ER values obtained with the BSCD and Betadex

columns. The average percent difference between the two columns was 3.6 ± 2.9 % for water and 3.0 ± 2.6 % for air

(41 pairs). Repeated injections of an α-HCH standard on both columns resulted in ER of 1.00 ± 0.01 for the BSCD and

0.99 ± 0.01 for the Betadex. The ion ratios 255/257 or 255/253 for samples were all within ± 5% of the standard ion

ratio.

HCHs in the Water Column

The physical structure of the water column changes on the transect from the Bering Sea through the Chukchi Sea

across the pole and southward through the Greenland Sea. In the continental shelf regions of the Bering-Chukchi seas,

the water column is vertically well mixed; thermoclines are temporarily formed and broken up in summer so the salinity

does not change markedly with depth (Hinckley et al., 1991). In the northern Chukchi Sea (station 7) the surface layer

consists of cold, low salinity water (-1.7oC, 30.5 psu) due to ice melt. This is underlaid by higher nutrient, more saline

water of Pacific origin (33.1 psu, 100-150 m). Below this is a layer of Atlantic water with higher salinity and

temperature (0.4-1.0oC, 34.5-34.8 psu, 200-800 m) (McLaughlin, et al., 1995). In the Canada Basin the cold, low

salinity polar mixed layer (~60 m) is underlaid by a broad halocline extending over several hundred meters. This

halocline is a permanent feature of the central Arctic Ocean. It is maintained by advection of sinking brine from the

freezing of ice on the continental shelves and prevents the surface layer from mixing with the deeper water (Hargrave et

al., 1988; Schlosser et al., 1995). This stratification does not exist in the Greenland Sea, where the water column is well

mixed (station 39).

Concentrations of α-HCH and γ-HCH in surface water and at depth from AOS-94 and BERPAC-93 are given in

Table 1 and are compared with other measurements in Table 2. Combined data from the two cruises are shown in Figure

2a. In the upper 40 m, α-HCH was 1.95 ± 0.26 ng L-1 in the southern Bering Sea and 2.08 ± 0.48 ng L-1 in the northern

Chukchi and averaged 2.02 ± 0.38 ng L-1 over the entire Bering-Chukchi region. The mean of the γ-HCH was 0.45 ±

0.09 ng L-1 and concentrations showed no trend with latitude. In the Bering and Chukchi seas α-HCH decreased slowly

to 1.5-1.7 ng L-1 at 300-350 m, whereas γ-HCH did not change significantly over this depth range. Concentrations were

two times lower in top 50 m of the Eurasia Basin (station 36-38) and the Greenland Sea (station 39): 0.87 ± 0.22 ng L-1

α-HCH and 0.20 ± 0.03 ng L-1 γ-HCH. At stations 37-38, just north of Spitzbergen, α-HCH was 0.95-1.44 ng L-1 at 10-

141

109 m, decreased to 0.53 ng L-1 at 235 m and 0.26 ng L-1 at 762 m; γ-HCH exhibited no significantly trend in the upper

100 m, although a decline was suggested for deeper samples. Samples taken in the Greenland Gyre (station 39) showed

little variation with depth, α-HCH ranged from 0.59-0.63 ng L-1 and γ-HCH from 0.17-0.20 ng L-1 at 10-540 m (Table

1).

Figure 2: Latitudinal trends of HCHs on AOS-94 and BERPAC-93. Bering Sea to the North Pole = increasing numbers; Pole to the Greenland Sea = decreasing numbers. a) α- and γ-HCH concentrations in water b) α- and γ-HCH concentrations in air c) α/γ-HCH ratio in air and water d) Fugacity ratios of α- and γ-HCH

Concentrations in the polar mixed layer (60 m) of the Arctic Ocean (station 7-35) averaged 2.42 ± 0.23 ng L-1

α-HCH and 0.47 ± 0.11 ng L-1 γ-HCH (Figure 2A). Concentrations in the central Arctic Ocean are higher than the

Atlantic side concentrations, but are not significantly different from values in the Bering and Chukchi Seas. At a depth

greater than 100m, HCHs decreased more sharply than in the regional seas. This may be attributed to the more stratified

water column structure discussed earlier. At 60-115 m the average concentrations were not significantly different than

0

20

40

60

80

100

120

140

50-5

455

-59

60-6

465

-69

70-7

475

-79

80-8

585

-89

89-8

584

-80

77-6

767

-61

55-4

50

0.5

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1.5

2

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a-HCHg-HCH

50-5

455

-59

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4

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8

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Water Air

0

0.5

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1.5

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2.5

a-HCHg-HCH

Air pg/m

Water (ng/L) α/γ−HCH Ratio

Fugacity Ratiob

c

d

Bering SeaGreenlandSea

3

a

Latitude N Latitude N

142

the surface samples and averaged 2.25 ± 0.24 ng L-1 α-HCH and 0.52 ± 0.18 ng L-1 γ-HCH, then dropped off rapidly to

0.62 ± 0.24 ng L-1 α-HCH and 0.26 ± 0.13 ng L-1 γ-HCH at 200-350 m. These are averages for stations 7-35, but the

vertical profile varied. At stations 7-8, α-HCH dropped from 2.49 ng L-1 at the surface to 1.40 ng L-1 α-HCH at 100 m,

but at more northern stations, the surface water like concentrations extended deeper in the water column. At station 24,

α-HCH was 2.49 ng L-1 at the surface and did not change substantially at 100 m (2.41 ng L-1) although by 250-300 m α-

HCH decreased by about five fold (0.49 ng L-1 ).

HCHs in the Air

Data for the air samples collected on AOS-94 are given in Table 3 and Figure 2. Concentrations of α-HCH in air

over the Bering and Chukchi seas (samples A-F) averaged 125 ± 17 pg m-3, compared to 91 ± 23 pg m-3 found on

BERPAC-93 (Jantunen and Bidleman, 1995), decreased to 64 ± 17 pg m-3 over the ice cap (samples G-U) and increased

again over the North Atlantic (sample V) to 131 pg m-3. Thus, the trend of α-HCH in air with latitude was opposite to

that in water (Figure 2B). The lower concentration of α-HCH over the central Arctic Ocean may result from wet

scavenging by fog and drizzle, which are predominant over the ice and inhibited volatilization from the surface water.

The levels of α-HCH over the ice cap are similar to those measured at the alert air monitoring station (82.5oN, 162oW)

on Ellesmere Island, Northwest Territories, Canada (Fellin et al., 1996). The AOS’94 Results confirm the 1992-1993

measurements in the Canadian Arctic (Falconer et al., 1995a) and the Bering and Chukchi Seas (Jantunen et al., 1995)

which show a threefold or greater decrease in α-HCH concentrations form the mid-1980s.

Ratios of α/γ in air on BERPAC-93 and AOS-94 ranged from 1.3-15 (Table 3 and Figure 2). The ratio of α-

HCH/γ-HCH in air has been suggested to follow the transport of technical HCH versus lindane (Hoff et al., 1992;

Oehme, 1991). From the composition of the technical mixture it is expected that the α/γ be 4-7 (Hinckley et al., 1991;

Iwata et al., 1993a). Lower ratios indicate sources of lindane superimposed on a technical HCH background. Oehme

(1991) and Pacyna and Oehme (1988) found that episodes of lindane transport from Europe were reflected in occasional

dips in the α/γ ratio at Spitzbergen. Higher proportions of α-HCH may indicate preferential removal of γ-HCH from air

by precipitation or gas-phase deposition to the sea surface during transport from source regions (Iwata et al., 1993b).

Thus the differences in the α/γ in air shown in Figure 2C are probably related more to transport pathways than to

latitudinal gradients.

Ratios of α/γ in surface water on AOS-94 and BERPAC-93 were within the range expected for technical HCH

(Table 1). The proportion of α-HCH increased from ~4.5 in the Bering Sea to ~6.5 at 75-79oN, then dropped to ~4.5

again in the northern Canada Basin and the Greenland Sea (Figure 2C). The α/γ ratio decreased with depth (Table 1)

reflecting preferential loss of α-HCH. The enantiomeric signatures of α-HCH provide direct evidence that HCHs are

subject to biological attack in the water column (see section on enantiomeric composition).

143

Table 1. Hydrographic Information, Concentrations o f αααα- and γγγγ-HCHs and Enantiomeric Ratio (ER) of αααα-HCH in Water

Latitude Temperature Salinity Concentration (ng L -1 ) ER of α α α α-HCH

Station (North) Longitude oC (psu) Depth (m) αααα-HCH γγγγ-HCH α/γα/γα/γα/γ Ratio BSCD Betadex_____________________________________________________________________________________________________________________AOS-94

AOS-1 67o47' 168o47' W -0.1 32.69 17 1.70 0.32 5.31 1.24 1.22

-0.3 34.74 47 1.19 0.34 3.50 1.09 1.09

AOS-2 72o08' 168o 50' W -1.7 32.14 10 1.29 0.25 5.16 1.11 1.09

-1.8 33.10 50 1.51 0.22 6.86 1.09 1.05

AOS-7 75o 00' 169o 59' W -1.5 30.63 12 2.49 0.41 6.07 0.97 1.02

AOS-8 75o 27' 170o 35' W -1.4 33.08 105 1.40 0.24 5.83 0.94 0.93

1.0 34.81 318 0.57 0.18 3.17 0.48 0.48

0.2 34.81 700 0.45 0.18 2.50 0.29 0.38

AOS-9 75o 45' 171o 14' W -0.4 34.94 1593 0.12 0.08 1.50 0.40 0.39

AOS-10 75o 57' 171o 40' W -0.4 34.95 1879 0.04 0.08 0.50 0.75 0.67

AOS-11 76o 38' 173o 19' W -1.6 31.01 10 2.36 0.37 6.38 1.06 1.04

-1.3 33.34 101 2.00 0.34 5.88 0.90 0.86

-0.4 34.95 2222 0.02 0.07 0.29 0.89 0.65

AOS-13 77o 48' 176o 18' W -1.6 30.44 10 2.55 0.36 7.08 0.91 0.89

AOS-16 78o 59' 175o 49' W -1.6 30.15 10 2.40 0.40 6.00 0.87 0.82

AOS-18 80o 09' 173o 15' W -1.7 32.43 69 2.16 0.33 6.55 0.89 0.92

-0.4 34.54 193 0.71 0.23 3.09 0.10 0.29

AOS-19 80o 09' 176o 46' W -1.6 30.73 10 2.18 0.40 5.45 0.81 0.83

AOS-21 80o 43' 179o 59' W -0.5 34.93 1613 0.10 0.12 0.83 0.67 0.54

AOS-24 82o 28' 175o 40' E -1.7 32.07 10 2.49 0.47 5.30 0.87 0.85

-1.5 34.16 100 2.41 0.46 5.24 0.91 0.93

0.5 34.81 303 0.49 0.20 2.45 0.20 0.15

AOS-25 83o 10' 173o 56' E -1.7 32.23 10 2.62 0.58 4.52 0.89 0.93

0.8 34.86 345 0.95 0.46 2.07 0.12 0.16

AOS-26 84o 04' 175o 04' E -1.6 31.05 10 2.07 0.58 3.57 0.90 0.93

AOS-28 85o 54' 166o 42' E -1.6 31.57 10 2.74 0.38 7.21 0.91 0.85

-1.3 34.28 113 2.59 0.74 3.50 0.89 0.86

144

Latitude Temperature Salinity Concentration (ng L -1 ) ER of α α α α-HCH

Station (North) Longitude oC (psu) Depth (m) αααα-HCH γγγγ-HCH α/γα/γα/γα/γ Ratio BSCD Betadex_____________________________________________________________________________________________________________________AOS-94

AOS-29 87o 09' 160o 42' E -1.7 32.10 10 2.09 0.70 2.99 0.84 0.88

AOS-30 88o 04' 174o 50' E -1.7 33.64 57 2.63 0.52 5.06 0.92 0.90

AOS-31 88o 47' 142o 44' E -1.7 32.58 10 2.43 0.56 4.34 0.93 0.90

1.1 34.88 330 0.4 0.15 2.67 0.13 0.08

AOS-35 90o 00' -1.7 31.89 10 2.69 0.55 4.89 0.87 0.88

-1.4 34.16 100 2.10 0.64 3.28 0.74 0.73

1.2 34.83 250 0.43 0.16 2.69 0.19 0.18

-1.0 34.94 4287 0.12 0.18 0.67 0.88 na b

AOS-37 84o 15' 35o 05' E -1.6 33.41 10 0.95 0.22 4.32 na 0.80

AOS-38 83o 51' 35o 41' E -1.8 34.09 47 1.04 0.17 6.12 0.82 0.79

-0.9 34.38 109 1.44 0.19 7.58 0.64 0.60

2.1 34.09 235 0.53 0.08 6.63 0.45 0.43

0.2 34.91 762 0.26 0.08 3.25 0.14 0.13

AOS-39 75o 00' 06o 03' W 4.0 34.29 10 0.63 0.20 3.15 0.75 0.67

0.5 34.89 75 0.59 0.20 2.95 0.60 0.59

-0.15 34.90 200 0.63 0.19 3.32 0.59 0.56

-0.97 34.86 540 0.61 0.17 3.59 0.55 0.54

____________________________________________________________________________________________________________________BERPAC-93 a

BC-South Bering 53-64 165W-176E 1-30 1.99 0.45 4.42 1.13 na

50-350 1.69 0.39 4.33 1.07 na

BC-Gulf of Anadyr 62-64 169-172W 1-35 1.94 0.47 4.13 1.09 na

50-80 1.69 0.41 4.12 1.14 na

BC-Chirikov Basin 62-65 174 W 1-40 2.08 0.47 4.42 1.13 na

BC-Alaska Chukchi 68-73 160-169W 1-40 2.16 0.45 4.80 1.06 na

53-106 1.69 0.49 3.45 0.85 na

133-298 0.73 0.22 3.32 0.76 na

BC-Siberian Chukchi 67-71 172W-179E 1-40 2.06 0.43 4.79 1.05 na

_____________________________________________________________________________________________________________________

a BC: 1993 (Jantunen and Bidleman, 1995), also see for temperature and salinity information.

b na: not analyzed

145

Table 2: HCH Concentrations in Surface Water in Sub -arctic and Arctic Regions

Concentration (ng L -1 )

Location Year αααα-HCH γγγγ-HCH Total Reference_____________________________________________________________________________________________

Canada Basin 1994 2.31 0.49 2.80 This work

1992 nr a nr 5.5 Macdonald et al., 1994.

1987 7.10 0.81 7.91 Patton et al., 1989

1986 4.47 0.61 5.08 Hargrave et al., 1988

Eurasia Basin 1994 0.95 0.22 1.17 This work

1985 1.03 0.22 1.25 Gaul, 1992

1979 1.40 0.32 1.72 Gaul, 1992

Greenland Sea 1994 0.65 0.18 0.83 This work

Russian Arctic 1990 nr nr 2.02 Melnikov and Vlasov, 1992 andMuir et al., 1992

Bering Sea 1993 1.95 0.46 2.41 Jantunen and Bidleman, 1995

1989-90 1.50 0.19 1.69 Iwata et al., 1993

1988 2.30 0.57 2.87 Hinckley et al., 1991

1979 nr nr 3.91 Tanabe and Tatsukawa, 1980

Chukchi Sea 1994 1.82 0.33 2.15 This work

1993 2.16 0.45 2.61 Jantunen and Bidleman, 1995

1989-90 1.40 0.18 1.58 Iwata et al., 1993

1988 2.30 0.61 2.91 Hinckley et al., 1991

N. Pacific and Bering Sea 1981 2.75 0.65 3.40 Kawano et al., 1988

Canadian Archipelago 1992 4.70 0.44 5.14 Falconer et al., 1995a(Resolute Bay, NWT)

North Pacific (Aleutians) 1987 2.80 0.59 3.39 Kurtz and Atlas, 1990

Gulf of Alaska 1989-90 1.60 0.26 1.86 Iwata et al., 1993_____________________________________________________________________________________________

a not reported

146

Table 3: Concentration of HCHs and Enantiomeric Ra tios (ERs) of αααα-HCH in Air

Latitude (N) Longitude Temperature Concentration (pg m -3 ) ER of αααα-HCH

Sample start/stop start/stop Air oC αααα-HCH γγγγ-HCH α/γα/γα/γα/γ Ratio BSCD Betadex

AOS-94

AOS A 49/50 132/137 W 20.0 155.4 13.5 11.51 1.13 1.20

AOS B 50/53 137/145 W 15.0 126.5 9.5 13.32 1.23 1.23

AOS C 53/53 145/157 W 10.3 112.6 7.4 15.22 1.20 1.15

AOS D 53/59 157/168 W 8.4 130.7 10.8 12.10 1.08 1.12

AOS E 59/64 168/166 W 8.2 107.4 47.3 2.27 1.14 1.15

AOS F 64/68 166/168 W 9.2 114.5 19.9 5.75 1.12 1.08

AOS G 69/70 169/169 W 7.6 30.1 6.1 4.93 1.03 1.12

AOS H 71/73 169/169 W 1.6 65.4 20.6 3.17 1.01 1.00

AOS I 73/74 168/169 W 1.0 53.2 12.1 4.40 1.07 1.08

AOS J 75/76 170/171 W -0.9 80.2 16.9 4.75 0.87 0.96

AOS K 76/77 171/175 W -1.0 64.0 16.5 3.88 1.07 1.01

AOS L 78/79 176/175 W -0.2 91.9 38.7 2.37 0.99 0.99

AOS M 80/81 173/178 W 0.2 34.1 26.8 1.27 1.07 1.02

AOS N 81/84 178/174 W 0.3 59.9 11.9 5.03 1.03 1.03

AOS O 84/87 173/159 W -0.3 59.9 7.4 8.09 1.02 1.03

AOS P 87/89 152/144 W -2.3 80.7 10.7 7.54 0.97 0.97

AOS Q 89/85 37/37 E -3.1 50.9 6.3 8.08 0.95 0.93

AOS R 85/84 37/35 E -2.8 76.3 10.3 7.41 0.96 0.97

AOS S 82/79 26/6 E -4.9 73.8 11.7 6.31 0.95 0.98

AOS T 77/67 0/23 W 1.8 70.1 9.6 7.30 0.88 0.91

AOS U 67/61 23/33 W 8.9 67.3 7.9 8.52 0.94 0.98

AOS V 55/45 51/61 W 12.6 131.5 14.5 9.07 0.91 0.93

BERPAC-93 a

BC-South Bering 105 23 4.57 1.06 na b

BC-Gulf of Anadyr 87 28 3.11 1.11 na

BC-Alaskan Chukchi 74 19 3.89 1.06 na

BC-Siberian Chukchi 70 20 3.50 1.03 na

a) See Jantunen and Bidleman, 1995 for sample location and air temperature.

b) na: not analyzed

147

Air-Water Gas Exchange of HCHs

The water/air fugacity ratio describes the saturation state of the water relative to the partial pressure of HCHs in

air. Fugacity ratios were calculated by (Bidleman and McConnell, 1995; Jantunen and Bidleman, 1995):

(1)

(2)

(3)

where, fw and fa are the fugacities in water and air (Pa), Cw and Ca are the concentrations of the dissolved and gaseous

HCHs in water and air (ng m-3), R is the gas constant (8.314 Pa m3 deg -1 mol -1 ), Ta is temperature of the air (K), H is

the Henry's law constant (Pa m3 mol -1) at the temperature of the surface water and M is the molecular weight (291 for

HCHs). A fugacity ratio = 1.0 implies that the HCHs in water and air are at equilibrium. The net flux is zero, although

HCHs are exchanged between air and water at the same rate. Fugacity ratios >1.0 and <1.0 indicate net volatilization

and deposition. HCHs in air and water from 1993 (Jantunen and Bidleman, 1995) and 1994 (Tables 1 and 3) were

averaged in 5o latitude bands to calculate the fugacity ratios. Henry's law constants for seawater at the average surface

water temperature were calculated according to Kucklick et al. (1991). Uncertainties in fugacity ratios were estimated

by propagation of errors in the air and water concentrations and the Henry's law constants as described previously

(Jantunen and Bidleman, 1995). Uncertainties were ±28% and ±23% for α-HCH and γ-HCH on BERPAC-93 (Jantunen

and Bidleman, 1995) and ±21% and ±22% on AOS-94.

Fugacity calculations show that all regions from the Bering and Chukchi seas to the Arctic Ocean are close to

equilibrium or oversaturated in α-HCH. Fugacity ratios are depositional in the Nansen Basin and Greenland Sea.

Fugacity ratios of γ-HCH indicate oversaturation in the Bering Sea and Nansen basin but undersaturated in the Chukchi

Sea, Greenland Sea and Canada Basin.

In the mid-1980s, the fluxes of HCHs were deposition in the Canada Basin at the Ice Island (81N,

RTC

HC = f

f

MRTC10 = f

M

HC10 = f

aa

w

a

w

aa9-

a

w-9

w

148

Table 4: Fugacity Ratio and Flux Calculations

αααα-HCH γγγγ-HCH

Water Air Location Average Temperature, K D-value a Fraction Open Windspeed H Flux ng m -2 d -1 H Flux ng m -2 d -1

Sample Sample Latitude (N) Water Air (mol m -2 d -1 Pa -1 ) Ocean (%) m s -1 Pa m3 mol -1 fw/fa Potential Actual b Pa m3 mol -1 fw/fa Potential Actual________________________________________________________________________________________________________________________________________________________________________________________________________________________

BC c BC, AOS 1-3 d 50-54 283.5 283.5 0.22 100 6.8 0.254 1.58 35.70 35.70 0.139 1.76 4.80 4.80

BC BC, AOS 4 55-59 283.3 281.5 0.30 100 9.2 0.250 1.95 77.30 77.30 0.137 1.45 6.35 6.35

BC BC, AOS 5-6 60-64 281.9 281.8 0.18 100 5.8 0.221 1.74 34.38 34.37 0.123 0.80 -2.53 -2.53

BC, AOS 1 e BC, AOS 7 65-69 277.4 280.7 0.34 90 10.2 0.150 2.06 51.47 46.32 0.086 1.02 0.26 0.26

BC, AOS 2 AOS 8-9 70-74 271.6 274.2 0.16 60 4.5 0.089 1.28 6.66 4.00 0.053 0.71 -1.44 -1.30

AOS 7-16 AOS 10-12 75-79 271.6 272.3 0.20 10 6.0 0.089 1.21 7.78 0.78 0.053 0.38 -6.58 -3.95

AOS 19-26 AOS 13-14 80-84 271.5 273.4 0.13 5 3.7 0.088 1.84 11.79 0.59 0.053 0.61 -2.18 -0.22

AOS 28-31 AOS 15-16 85-89 271.4 271.8 0.23 5 6.8 0.087 1.33 11.92 0.60 0.053 1.41 1.83 0.09

AOS 35 AOS 17 89-85 271.4 270.1 0.15 5 4.4 0.087 2.06 18.46 0.92 0.053 1.94 2.10 0.11

AOS 37 AOS 18-19 84-80 271.5 269.9 0.10 50 2.7 0.088 0.48 -9.10 -4.55 0.053 0.47 -1.32 -0.66

AOS 39 AOS 20 77-67 277.2 275.0 0.27 100 8.1 0.147 0.60 -17.22 -17.22 0.085 0.67 -1.86 -1.86__________________________________________________________________________________________________________________________________________________________________________________________________________________________

a) D = 86400 ka/RT (mol m -2 d-1 Pa -1 ), where the air-side mass transfer coefficient ka (m s-1 ) is calculated as a function of wind speed (McConnell et al., 1993).

b) actual flux = potential flux x fraction of open ocean.

c) BC - average of water samples collected in this region on BERPAC-93 (Jantunen and Bidleman, 1995).

d) BC, AOS 1-3, etc. = average of air samples collected in this region on the two cruises.

e) BC, AOS 1, etc. = average of surface water samples collected in this region on the two cruises.

149

100W) (Cotham and Bidleman, 1991) and the Bering and Chukchi seas (Hinckley et al., 1991). The driving force for the

reversal in 1993-94 is the decrease in atmospheric concentrations of HCHs, while HCHs in water have remained

essential constant (Bidleman et al., 1995b; Jantunen and Bidleman, 1995).

Net fluxes were calculated from the two film model with fugacity definitions using mass transfer coefficients as a

function of wind speed (Bidleman and McConnell, 1995; Jantunen and Bidleman, 1995; McConnell et al., 1993) (Table

4).

(4)

(5)

Since the exchange of HCHs is gas-phase controlled (Cotham and Bidleman, 1991; Hinckley et al., 1991), the mass

transfer coefficient for air (ka, m s-1 ) is used in equation 5. Fluxes of α- and γ-HCH at different northern latitudes are

summarized in Figure 3A and 3B. Positive and negative values indicate volatilization and deposition. Fluxes are shown

for 100% open water (potential flux) and adjusted for fractional ice cover (actual flux). Fractional ice cover values were

taken from microwave satellite data supplied by the US Coast Guard and ice maps provided by long range aircraft.

Potential and actual fluxes coincide in the Bering Sea and the Greenland Sea, which are completely unfrozen during

August to September. Even though fugacity ratios predict that potential fluxes in the Canada and Eurasia basins are high

(Figure 2D) the actual fluxes are much lower because of ice cover (Figure 3A and 3B). In the high Arctic this severely

limits the air-water exchange of HCHs.

Enantiomeric Composition of α-HCH in the Water Column

Enantiomer ratios (ER) of α-HCH in water are given in Table 1 and surface water values are averaged by 5o

latitudinal bands in Figure 4. Samples from the Canada Basin and Greenland Sea were depleted in the (+)α-HCH,

resulting in ER values <1.00. An opposite enantiomer degradation pattern was found in the Bering and Chukchi seas,

where the ERs in surface water were generally >1.00 indicating selective breakdown of (-) α-HCH. Reasons for the

opposite enantioselectivity are not known, but may be related to different biodegradation pathways in the water types.

Studies done in the North Sea also show reversals of the ERs in different regions (Faller et al., 1991a). Falconer et al.

(1995a,b) found degradation of (+) α-HCH in the Canadian Archipelago and nearby Amituk Lake. Buser and Müller

(1995) noted that (+)α-HCH was more rapidly degraded in sewage sludge. In contrast, selective degradation of (-)α-

HCH has been noted in soils from Germany (Müller et al., 1992) and western Canada (Falconer et al., 1995c).

RTk 86400

= )Pa d m (molD

)f-f(D M 10 = )d m F(ng

a1-1-2-aw

awaw9-1-2

150

Figure 3: Potential and actual net fluxes of α- and γ-HCH at different latitudes. Actual flux = potential flux x fraction of open water. Positive flux = volatilization, negative flux = deposition.

-40

-20

0

20

40

60

80

100

50-5

4

55-5

9

60-6

4

65-6

9

70-7

4

75-7

9

80-8

4

85-8

9

89-8

5

84-8

0

77-6

7

Latitude (N)

Flu

x (n

g m

-2 d

ay-1

)

-8

-6

-4

-2

0

2

4

6

8

50-5

4

55-5

9

60-6

4

65-6

9

70-7

4

75-7

9

80-8

4

85-8

9

89-8

5

84-8

0

77-6

7

Latitude (N)

Flu

x (n

g m

-2 d

ay-1

)

151

In the Canada Basin and Greenland Sea preferential loss of (+)α-HCH increased with depth (Table 1). Figure 5

shows a profile from north of Spitzbergen (station 38) where the ER fell from 0.82 at 47 m to 0.14 at 753 m. Similar

losses of (+)α-HCH were seen in water from 200-350 m at stations 18, 24, 25, 31 and 35. A few stations showed less

change in enantioselectivity with depth. ERs at station 11 were 1.05 at 10 m, 0.93 at 100 m and 0.65-0.71 at 2222 m.

Difference among these stations may reflect the physical structure and microbial activity in the water column. In

general, little difference was seen in ER values of surface and deep water samples from the Bering and Chukchi seas

except for the most northern stations in the Chukchi Sea (Station BC 48-51) where the ERs were 1.07 at 1-40 m, 0.85 at

53-106 m and 0.76 at 133-298 m.

Figure 4: Enantiomeric ratios (ERs) of α-HCH in air and water at different latitudes. ER = (+)α-HCH/(-)α-HCH.

Enantiomers as Tracers of Air-Water Gas Exchange

The α-HCH in air samples taken in 1993 and 1994 within 40 m of the sea surface was non-racemic and followed

the same order of degradation as in surface water. Air over the Bering Sea and southern Chukchi Sea was depleted in (-

)α-HCH (ER>1.00) but, as with the water, the selectivity reversed at higher latitudes. ERs in air over portions of the

Arctic Ocean and the northern Atlantic Ocean were <1.00, from depletion of (+)α-HCH. The similarities in air and

water enantiomeric profiles are shown in Figure 4 for the Bering Sea and the Greenland Sea. These suggest that sea-to-

air gas exchange is an important source of α-HCH to the marine boundary layer.

50-54 55-59 60-64 65-69 70-74 75-79 80-84 85-89 89-85 84-80 77-67 67-61 55-450.7

0.8

0.9

1.0

1.1

1.2

Water Air

Latitude N

ER

50-54 55-59 60-64 65-69 70-74 75-79 80-84 85-89 89-85 84-80 77-67 67-61 55-450.7

0.8

0.9

1.0

1.1

1.2

Water Air

Latitude N

ER

152

Air-water gas exchange is a "two way street". At equilibrium the net flux is zero but volatilization and deposition

still occur at the equal rates (Hornbuckle and Eisenreich, 1995; Murphy, 1995). This concept has been applied to the

exchange of enantiomers (which have the same Henry's law constant) when the α-HCH in the bulk air is racemic and the

α-HCH in surface water is non-racemic (Ridal et al., 1996). The concentration of α-HCH in seawater is ~104 greater

than in air. Because of this buffering capacity the ER of α-HCH in air will tend toward the seawater value regardless of

whether air-water equilibrium is approached from the deposition or volatilization side. Thus a non-racemic ER value for

air does not in itself indicate net volatilization of α-HCH, although this is implied from the fugacity ratios (Figure 2D).

The proportion of α-HCH in the atmosphere that has volatilized from the ocean can be estimated from the ER values in

boundary-layer air (ERair) and surface water (ERwater) assuming that: the ER of volatilized α-HCH is equal to ERwater (≠

1.00) and the α-HCH in bulk air is racemic (ERbg=1.00). The fraction of α-HCH in the air column arising from

outgassing is:

(6)

Thus for ERmix = 0.90 and ERwater = 0.80, f = 0.47.

The α-HCH in air over the Canada Basin north of 75oN is racemic even when the surface water is not. This is

possibly due to inhibition of sea-to-air gas exchange by the ice cover, even though the fugacity ratios predict that the

water is oversaturated. Figures 3A and 3B shows a large difference between the potential and actual fluxes in the high

Arctic. Also, mixing of air masses from the east and west that are depleted in the opposite enantiomer may account for

the racemic α-HCH over the northern Canada Basin.

Five day back trajectories were calculated for each air sample at 925mb and 850mb. Air masses over the northern

Chukchi (Sample K-P) were transported from northern Russia and the Bering Strait, whereas those on the eastern side of

the North Pole (samples Q-S) arrived from the North from the North Atlantic and Scandinavian regions.

Figure 5: Chromatograms (BSCD column) showing enantioselective degradation of α-HCH with depth at stations AOS-37 and 38.

1))(ERER-(ER

1))(ERER-(ER=

mixbgwater

waterbgmixwater +

+f

109 m 235 m 753 m

α-HCH Standard 10 m 47 m

(+)(-)

109 m 235 m 753 m

α-HCH Standard 10 m 47 m

(+)(-)

153

Conclusions

Air samples taken in arctic and subarctic regions in 1993 and 1994, show a three- to five-fold decrease in α-HCH

since the mid-1980s apparently due to global restrictions in usage of technical HCH products. During summer,

atmospheric levels of α-HCH over the polar ice cap are about a half those over the regional seas. Concentrations of

HCH in surface water remained constant with latitude west of the North Pole, being approximately the same under the

polar ice cap as in the Bering and Chukchi seas but twice as high in the Greenland Sea. HCHs in surface water of the

Bering and Chukchi seas have not changed substantially since the mid-1980s. A different trend was seen for α-HCH in

air, which was reduced over the polar cap by about 25-45% compared to levels over the regional seas. Peak

concentrations under the polar cap are two to three times higher than in regional seas. HCHs in surface water of the

Bering and Chukchi seas have not changed substantially since the mid-1980s.

The decrease in atmospheric HCHs over the northern oceans has caused a reversal in the air-sea flux direction,

from net deposition in the 1980s to net volatilization in the 1990s. The potential for volatilization is greatest in the

Canada Basin, where surface water concentrations are highest and atmospheric concentrations are minimal but the actual

amount of degassing from the surface layer is limited by the small unfrozen area. Open leads, melt holes and polynyas

are probably important sites of volatilization to the atmosphere in the high Arctic.

Enantioselective degradation of α-HCH takes place in arctic and subarctic waters, and the selectivity is reversed in

different regions. (-)α-HCH is depleted in the Bering and Chukchi seas, whereas (+)α-HCH is preferentially lost in the

Arctic Ocean and the Greenland Sea. Chiral degradation of α-HCH increases with depth in the Arctic Ocean to the point

where the (+) enantiomer is almost absent in deep samples. Enantiomeric profiles of α-HCH in air and surface water

show similar patterns implying that revolatilization of α-HCH from the ocean surface contributes to α-HCH in the

atmospheric boundary layer.

Acknowledgements

This work was supported by the Canadian Department of Indian and Northern Affairs, North Contaminants Program.

We thank the Russian Committee for Hydrometeorology for ship time and the U.S. Fish and Wildlife Service for logistic

support. Thanks also to Rob Macdonald and co-workers at the Institute for Ocean Sciences, Sidney, British Columbia,

for the use of their sampling equipment, Jim Swift at Scripps Institute of Oceanography, La Jolla, California for the

hydrographic information and Crispin Halsall (MSC) for the trajectory data.

154

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Paper 6

Hexachlorocyclohexanes (HCHs) in the Canadian Archipelago:

2. Air-Water Gas Exchange of αααα- and γγγγ-HCH

Liisa. M. Jantunen1, Paul A. Helm2, Henrik Kylin3, Terry F. Bidleman1,

Environmental Science and Technology, 2008, 42, 465-470.

1 Centre for Atmospheric Research Experiments, Environment Canada, 6248 Eighth Line, Egbert ON, L0L 1N0, Canada. 2 Chemical Engineering and Applied Chemistry, University of Toronto, 200 College St., Toronto, Ontario, M5S 3E5 Canada Current Address: Environmental Monitoring and Reporting Branch, Ontario Ministry of the Environment, 125 Resources Road, West Wing, Toronto, Ontario, M9P 3V6 Canada 3 Norwegian Institute for Air Research, Polar Environmental Centre, No-9296 Tromsø, Norway; Department of Environmental Assessment, Swedish University of Agricultural Sciences, P.O. Box 7050, SE-705-07, Uppsala, Sweden. Contributions: Liisa Jantunen and Paul Helm collected, processed and analysed samples from Resolute Bay and TNW’99. Liisa collected samples on the first leg of TNW’99; while Henrik Kylin collected, processed and analysed samples on a later leg of TNW’99. Liisa analysed the data and wrote the paper. Terry Bidleman secured funding and provided scientific guidance during every step of this project.

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Abstract

Air and water were sampled in the Canadian Archipelago during summer on the Tundra Northwest 1999

(TNW-99) expedition and air was sampled at Resolute Bay (RB), Nunavut, to determine the gas exchange of α- and

γ-hexachlorocyclohexanes (HCHs) and the enantiomers of α-HCH. Air concentrations of ΣHCH during TNW-99

and at RB were similar, averaging 55 and 53 pg m-3, respectively. The net gas exchange direction was volatilization

for α-HCH and near equilibrium or deposition for γ-HCH, while actual fluxes depended on the fraction of open

water. EFs in air sampled from shipboard were significantly correlated to those in surface water for events with

>90% open water, but were closer to racemic and not correlated to EFs in water for events with 0-50% open water.

Levels of α-HCH in air at RB averaged 37 ± 9 pg m-3 from June - early July and EFs were close to racemic (0.496 ±

0.003). In mid-July the ice pack broke up around RB. From this point through August, air concentrations increased

significantly to 53 ± 5 pg m-3 and the mean EF decreased significantly to 0.482 ± 0.010. Air concentrations of γ-

HCH at RB did not differ significantly before (8.0 ± 3.8 pg m-3) and after (6.6 ± 0.76 pg m-3) ice breakup. Results

show that α-HCH enantiomers are sensitive tracers for following the impact of ice cover loss on gas exchange in the

Arctic.

Introduction

Technical hexachlorocyclohexane (HCH) is a mixture of several isomers, the most abundant being the α-

HCH (60-70%), β-HCH (5-12%) and γ-HCH (10-15%) (1). Although γ-HCH is the only isomer with insecticidal

properties and β-HCH is generally the most bioaccumulating (2), all three HCHs are found in arctic biota.

Differences in the relative body burdens of the HCH isomers among species result from selective metabolism and

varying concentration distributions in Arctic Ocean water masses (3). HCH is the most abundant organochlorine

pesticide in arctic air (4) and water (5). The transport and mass balance of HCHs in the Arctic have been discussed

in recent reports (3,6,7). Technical HCH was heavily used in Asian countries until it was banned or heavily

restricted by China, the former Soviet Union and India between the mid-1980s and 1990 (8,9). Concentrations

of α-HCH in arctic air responded quickly to these large-scale usage changes and declined by an order of

magnitude from the early 1980s to mid-1990s in steps that closely matched global usage (8) and emission

estimates (9). As a consequence, the direction of net gas exchange in arctic waters reversed from deposition in

the 1980s to air-water equilibrium or volatilization in the mid-1990s (10-13).

Although HCHs have been reported in arctic air and water in the Bering-Chukchi seas (10,11), the northern

Canada Basin (11), the Eurasian Arctic Ocean (12,14), and in air at arctic monitoring stations (15), few studies have

been done in the main portion of the Canadian Archipelago. HCH concentrations in central Archipelago water were

measured at Resolute Bay on Cornwallis Island, NV (RB, 74.68oN, 94.90oW) in 1992 (16) and seasonal air-water

exchange at RB was followed in 1993 (17), using air data from the monitoring station at Alert, NV (82.50oN,

62.33oW). Air concentrations of HCHs were reported at Kinngait, NV (64.22oN, 76.53oW) in the eastern

Archipelago for 2000-2001, but no co-located water measurements were made (15). An assessment of surface water

concentrations and air-water exchange in the Arctic Ocean and adjoining seas was recently made for γ-HCH (18).

160

The purpose of this investigation was to determine the spatial distribution and air-water gas exchange of

HCHs across the Canadian Archipelago. Parallel water and air samples were collected in the Archipelago on a

Swedish-Canadian expedition, Tundra Northwest 1999 (TNW-99), with additional air samples collected at RB.

Spatial variability and pathways of HCHs in surface water are discussed in a separate paper (19). Here we report air-

sea exchange of α-HCH and γ-HCH as affected by seasonal ice cover and the use of α-HCH enantiomers to follow

the exchange.

Experimental Section


The TNW-99 expedition platform was on board the CCGS Louis S. St-Laurent. The ship traveled from

Nuuk, Greenland across the Davis Strait to Iqaluit, Canada then traversed the Archipelago from Hudson Strait to RB

(Leg 1), along a southern route to Tuktoyaktuk and the southern Beaufort Sea (no samples taken), and returned along

a northern route to RB, Ellef Ringnes Island, over Devon Island, through Baffin Bay and the Davis Strait (Leg 2). A

map of the cruise track is shown in Figure 1. Air and surface water samples on Leg 1 between Iqaluit and Resolute

Bay were collected by Environment Canada (EC) personnel while those on Leg 2 were collected by the Swedish

University for Agricultural Sciences (SLU). Additional air samples were collected at RB from June 7 to August 14,

1999. Collection and processing methods for air samples are given below. Dates and locations are summarized in

Table 1 and in detail in Table 2. Collection of water samples is described in (19).

Figure 1: Map of TNW-99 Cruise Track.

Baffin BayBeaufort Sea

Hudson Bay

Resolute Bay


Hudson Bay

Resolute Bay


Hudson Bay

Resolute Bay

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Table 1: Atmospheric concentrations of HCHs during TNW-99 and at Resolute Bay (RB)a, pg m-3

αααα-HCH γγγγ-HCH

pg m-3 pg m-3 αααα-HCHDate (1999) mean s.d. mean s.d. αααα-HCH/ γγγγ-HCH EF s.d.

Leg 1 TNW-99 June 30 - July 11 42 9.0 11 2 4.0 0.475 0.015

Leg 2 TNW-99 August 1 - 30 48 14 8.8 4.7 6.3 0.460 0.019

RB, before ice break up June 7 - July 17 37 9.2 8.0 3.7 5.1 0.496 0.0035

RB, after ice break up July 19 - Aug 14 53 5.0 6.6 0.76 8.2 0.483 0.0088

a: Details are given in Supporting Information S1 and S3.

EC: High volume air samples were continuously collected on the bow of the ship. When possible, the bow

was into the wind to avoid contamination. At RB, the air sampling site was located inland from Lancaster Sound, 6

km from the village of Resolute and 3 km from the local airport behind a ridge of hills. Air samples were collected

using sampling trains consisting of a glass fiber filter (EPM 2000 20 x 25 cm Whatman, Ltd., Maidstone, England)

followed by two polyurethane foam (PUF) plugs (8.0 cm diameter x 7.5 cm). Air volumes ranged from 462-689 m3

and 846-1652 m3 for TNW-99 and RB, respectively. Air sampling media were prepared and processed after sample

collection as described in (20). Extract volumes were brought to 1 mL (quantitative analysis) or 100 µL (enantiomer

analysis) through volume reduction under a nitrogen stream and solvent exchange into isooctane.

SLU: Air sampling was done on the deck above the bridge using the same sampling apparatus and PUF size

as EC but with a 15 cm diameter filter. Air volumes ranged from 520-710 m3. PUFs were Soxhlet extracted in

acetone. Air sample extracts were cleaned up on a gel permeation column with Biobeads SX-3, eluted with1:1

DCM:ethyl acetate to remove some of the larger molecules and then fractionated on deactivated silica (21).

Analysis was done using capillary GC-electron capture negative ion mass spectrometry. A DB-5

column was used for quantitative work and chiral-phase columns for analysis of α-HCH enantiomers, details are

provided in (19).


Quality Control

EC and SLU Air samples: Field blanks were done by placing a clean PUF and filter into the air

sampling apparatus and drawing air for 30 s. No peaks were detected in the blank PUFs; blank filters were not

analyzed. Each PUF was spiked with α-HCH-d6 prior to extraction, recoveries averaged, 80 ± 15%, (n=60, EC) and

92 ± 12% (n=15, SLU). Sample concentrations have been adjusted for mean recoveries. The saturation state of the

PUF was checked by analyzing the front and back PUFs separately. About half of the samples showed a low level of

breakthrough for α-HCH, ranging from 0.5-4%, whereas no breakthrough was seen for γ-HCH. Enantiomer

fractions of α-HCH, EF = areas of (+)/[(+) + (–)] enantiomers, were determined on two dissimilar chiral-phase

columns for EC samples on TNW-99 (n=9) and at RB (n=19), the average difference in EFs was 0.32%. A single

chiral column was used by SLU (19). Quality control for quantitative and enantiomer analyses of water samples is

162

reported in (19).

Table 2: Air and water concentrations of αααα- and γγγγ-HCH on TNW-99

Air/Water Latitude (N) Longitude (W) Air (pg m-3) Water (ng L-1)

Samplesa Date, 1999 on off on off αααα-HCH γγγγ-HCH αααα-HCH/ γγγγ-HCH αααα-HCH EF αααα-HCH γγγγ-HCH αααα-HCH/ γγγγ-HCH αααα-HCH EF

1/3-4 June 30-July1 62.0 62.5 67.7 73.5 50 13 3.8 0.469 2.40 0.27 8.89 0.459

2/3-4 July 1-2 62.5 62.5 73.5 73.6 59 12 4.9 0.473 2.40 0.27 8.89 0.457

3/5 July 2-3 62.5 64.1 73.6 79.0 45 10 4.5 0.469 2.30 0.24 9.58 0.451

4/5-6 July 3-4 64.1 66.4 79.0 80.9 40 11 3.6 0.477 2.40 0.23 10.43 0.451

5/6 July 4-5 66.4 67.6 80.9 81.2 42 12 3.6 0.469 2.50 0.22 11.36 0.441

6/6 July 5-6 67.6 67.7 81.2 81.4 45 13 3.5 inter 2.50 0.22 11.36 0.441

7/6 July 6-7 67.7 69.6 81.4 81.6 43 8.8 4.9 0.446 2.50 0.22 11.36 0.441

8/6 July 7-8 69.6 70.2 81.6 88.7 23 6.8 3.4 0.492 2.50 0.22 11.36 0.441

9/7 July 8-10 70.2 72.7 88.7 93.0 35 9.0 3.9 0.491 3.10 0.29 10.69 0.447

10/8-9 July 10-11 72.7 72.7 93.0 92.3 43 10 4.2 0.492 3.15 0.37 8.51 0.446

11/10-11 August 1-2 69.8 70.5 133.3 139.1 60 16 3.8 0.443 4.80 0.35 13.71 0.435

12/11-12 August 2-3 70.5 69.6 139.1 139.5 69 12 5.6 0.442 5.15 0.35 14.71 0.433

13/13 August 4-5 72.0 69.4 139.5 138.9 73 19 3.9 0.445 5.10 0.28 18.21 0.435

14/14-16 August 8-9 71.5 74.3 127.9 117.5 54 10 5.4 0.468 4.57 0.32 14.42 0.437

15/16 August 10-11 73.7 73.7 115.5 115.5 62 14 4.6 0.501 5.40 0.41 13.17 0.433

16/17 August 12-13 73.7 74.8 115.5 108.4 42 8.7 4.9 0.493 4.60 0.45 10.22 0.436

17/18-19 August 15-16 74.9 74.6 107.8 94.9 37 5.6 6.6 0.483 4.05 0.40 10.25 0.439

18/19 August 16-17 74.6 76.4 93.4 97.2 42 4.2 9.9 0.456 3.40 0.38 8.95 0.441

19/20-21 August 17-18 76.6 79.0 97.6 105.1 29 2.8 10.5 0.447 3.70 0.38 9.74 0.440

20/22 August 19-20 79.0 79.0 104.7 104.7 56 9.8 5.7 0.440 4.50 0.38 11.84 0.437

21/22 August 20-21 79.0 77.2 105.0 92.1 51 10 5.0 0.448 4.40 0.35 12.57 0.439

22/23 August 21-22 77.0 76.3 89.8 87.3 25 3.6 6.9 0.461 4.20 0.32 13.13 0.440

23/25 August 24-25 76.2 74.5 86.9 82.5 44 5.9 7.5 0.451 2.60 0.23 11.30 0.454

24/29 August 27-28 74.3 71.8 82.1 72.2 48 6.8 7.1 0.455 2.70 0.25 10.80 0.449

25/29 August 29-30 71.5 68.6 69.4 66.1 31 4.4 7.1 0.461 2.70 0.22 12.27 0.453

mean 46 10 5.4S.D. 13 3.9 1.9

a Air samples paired with water samples from ref 19.

HCHs in Air

Air concentrations of α-HCH from TNW-99 averaged 46 ± 13 pg m-3 (n=24) and compared very well with

those at RB, 44 ± 11 pg m-3 (n=19). For γ-HCH, results from TNW-99 averaged 9.5 ± 3.9 pg m-3, which is not

significantly different from the RB mean of 7.4 ± 2.9 pg m-3 (p >0.05) (Tables 1-3). Concentrations of α-HCH and

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γ-HCH were correlated on TNW-99 (r2 = 0.62, p<0.001). The relationship at RB was not significant (p = 0.10) and

skewed by four samples that were elevated for γ-HCH but not for α-HCH (Table 3: samples 1, 2, 9 and 10). When

these four samples were removed, the correlation was significant with r2 = 0.48 (p = 0.004).

Annual mean concentrations of α-HCH at Alert and Kinngait during 2000-2001 were 22 and 28 pg m-3,

while those of γ-HCH were 5.6 and 4.1 pg m-3, respectively (15). Slightly higher annual means were found at Alert

in 1999, 34 and 5.8 pg m-3 for α-HCH and γ-HCH (Su, Y. and Hung, H., Environment Canada, personal

communication). Means for α-HCH and γ-HCH at western arctic and subarctic stations in 2001-2002 were 19 and

2.7 pg m-3 at Point Barrow, AK (71.30oN, 156.60oW), and 48 and 4.5 pg m-3 at Little Fox Lake, YK (61.35oN,

135.63oW) (15). Shen et al. (22) deployed passive samplers throughout North America, including the Canadian

Arctic, during 2000-2001. Concentrations in the Arctic ranged from 34-95 pg m-3 for α-HCH and 5.3-15 pg m-3 for

γ-HCH, with the highest levels on the north shore of Baffin Island and the lowest at Alert. The concentrations found

on TNW-99 and at RB fall within the ranges in these studies. Hung et al. (4) found that α-HCH and γ-HCH declined

in air at Alert with times for 50% reduction (half-lives) of 9.1 and 5.7 y between 1993-1999. A slower decline of α-

HCH in arctic air was seen compared to stations on the Great Lakes, where half-times between 1996-2003 ranged

from 1.6 - 4.2 y (23). Half-lives for γ-HCH at Great Lakes stations ranged from 4.2-10 y (23), spanning the rate at

Alert. The slower decline of α-HCH in the Arctic may be due to regional buffering of air concentrations by

revolatilization during times of open water, see below.

Air-Water Gas Exchange

Fugacity ratios

Water samples collected during TNW-99 allowed estimates of air-water exchange to be made. Concentrations

of α-HCH and γ-HCH in surface water of the Archipelago are given in Table 2 (19). Examination of spatial trends

showed higher concentrations of HCHs in the southern Beaufort Sea, western and northern Archipelago, which

decreased to the east and south of Barrow Strait near RB. Enantiomer fractions of α-HCH, EF = areas of (+)/[(+) +

(–)] enantiomers, averaged 0.438 west and north of Barrow Strait and increased to an mean of 0.452 in the eastern

Archipelago. A plausible explanation for these trends is dilution of water advected from west of Barrow Strait with

Arctic Ocean water entering northern Baffin Bay via the Nares Strait and Atlantic Ocean water entering southern

Baffin Bay through the Davis Strait (19).

The water/air fugacity ratios (FR) of α-HCH and γ-HCH were calculated from:

AA

W

RTC

HCFR = (1)

where CW and CA are water and air concentrations (mol m-3), H is the Henry's law constant (Pa m3 mol-1) at the water

temperature, TA is the air temperature (K) and R is the gas constant (8.314 Pa m3 mol-1 K-1) (13,20). FR =1.0, >1.0

and <1.0 indicate steady state, net volatilization and net deposition, respectively. Freshwater Henry’s Law constants

were adjusted for temperature (13) and were assumed to be 20% higher in seawater due to the salting out effect

(13,24). The pooled relative standard deviation for replicate water sample analyses was 5.2% for α-HCH and 18%

164

for γ-HCH (19). Only single air samples were taken, so the precision was estimated from recoveries of the α-HCH-

d6 surrogate from spiked PUF plugs, ~15%. The Henry's law constants was 18% for α-HCH and 21% for γ-HCH

(13). The propagated uncertainty in FR, assuming uncertainties in CW, CA and H, is:

2222HCaCwFR RSDRSDRSDRSD ++= (2)

Which results in 24% for α-HCH and 32% for γ-HCH.

Fugacity ratios were calculated from TNW-99 parallel air and water samples, and at RB where air samples

were paired with water samples collected in the central Archipelago (stations 7,8,9,19 and 20, see Figure 1 and Table

1 in 19). In Table 4, the FRs are summarized into five areas with characteristic water concentrations (Table 1 in 19),

details are given in Table 5.

Fugacity ratios from TNW-99 ranged from 1.1-4.6 for α-HCH and 0.15-1.4 for γ-HCH (Table 5). Thus, α-

HCH was near equilibrium or undergoing net volatilization from the surface water, while γ-HCH was near

equilibrium or undergoing net deposition. At RB (Tables 4-5), the FR calculations showed the same trend, net

volatilization of α-HCH (average 2.1 ± 0.66, significantly >1.0 at p <0.001) and net deposition of γ-HCH (average

0.55 ± 0.18, significantly <1.0 at p < 0.001).

Two other studies of gas exchange at RB in 1992-1993 found similar results, net volatilization of α-HCH

and near equilibrium for γ-HCH (16,17). Fugacity ratios of α-HCH in the Bering and Beaufort seas in 1993-1994

were significantly greater than 1.0, while α-HCH in the Chukchi Sea and northern Canada Basin was near air-water

equilibrium. Net deposition of α-HCH was estimated in these four water bodies (13). Su et al. (15) estimated air-

water exchange of α-HCH using regionally varying water concentrations and a circumpolar average α-HCH in air of

23 pg m-3. Fugacity ratios ranged from 0.57 in the eastern Arctic Ocean to 6.7 in the central and western

Archipelago. Paired air and water measurements in the eastern Arctic Ocean in 1996 showed near-equilibrium

conditions for α-HCH and net deposition of γ-HCH (12). Weber et al. (18) estimated net deposition of γ-HCH over

most of the western Arctic Ocean and deposition or near-equilibrium in the central and eastern Arctic Ocean.

165

Table 3: Atmospheric concentrations and fluxes of αααα- and γγγγ-HCH at Resolute Baya.

Wind Air Open αααα-HCH γγγγ-HCH

Speed Temp αααα-HCH γγγγ-HCH αααα-HCH/ Water fW/fA b kA

c Actual flux (ng m-2 d-1)e Actual flux (ng m-2 d-1)

Date (m s-1) oC pg m-3 pg m-3 γγγγ-HCH αααα-HCH EF % αααα-HCH γγγγ-HCH (m s-1) DAW d ND NV N ND NV N

RB 1 June 7-9 4.8 -2 44 12 3.8 0.498 0 2.0 0.32 0.0043 0.166 0.0 0.0 0.0 0.0 0.0 0.0

RB 2 June 10-12 9.5 -3 46 12 3.7 0.495 0 1.9 0.31 0.0085 0.327 0.0 0.0 0.0 0.0 0.0 0.0

RB 3 June 14-16 3.4 1 33 6.0 5.6 0.496 0 2.5 0.62 0.0032 0.122 0.0 0.0 0.0 0.0 0.0 0.0

RB 4 June 17-19 10.3 0 24 4.0 5.9 0.501 0 3.6 0.91 0.0093 0.354 0.0 0.0 0.0 0.0 0.0 0.0

RB 5 June 21-23a 8.9 1 36 5.6 6.4 0.497 0 2.4 0.65 0.0079 0.298 0.0 0.0 0.0 0.0 0.0 0.0

RB 6 June 24-26 4.4 1 26 4.8 5.3 0.499 0 3.3 0.76 0.0040 0.151 0.0 0.0 0.0 0.0 0.0 0.0

RB 8 July 1-3 6.8 4 36 7.3 4.9 0.489 0 2.4 0.50 0.0060 0.224 0.0 0.0 0.0 0.0 0.0 0.0

RB 9 July 5-7 3.0 6 47 11 4.2 0.498 0 1.8 0.32 0.0029 0.110 0.0 0.0 0.0 0.0 0.0 0.0

RB 10 July 8-10 1.9 7 51 15 3.4 0.495 0 1.6 0.24 0.0022 0.080 0.0 0.0 0.0 0.0 0.0 0.0

RB 11 July 13-15 6.3 3 38 6.0 6.4 0.491 0 2.2 0.61 0.0056 0.209 0.0 0.0 0.0 0.0 0.0 0.0

RB 12 July 15-17 8.7 2 27 4.4 6.2 0.495 0 3.1 0.83 0.0077 0.291 0.0 0.0 0.0 0.0 0.0 0.0

RB 13 July 19-21 12.8 3 51 5.3 9.6 0.468 15 1.6 0.69 0.0119 0.449 -7.9 -13 -5.2 0.8 0.6 0.3

RB 14 July 22-24 8.5 4 59 7.1 8.4 0.472 20 1.4 0.52 0.0075 0.283 -7.7 -11 -3.3 0.9 0.5 0.4

RB 15 July 26-28 3.0 5 54 6.4 8.4 0.486 25 1.6 0.57 0.0030 0.110 -3.4 -5.4 -2.0 0.4 0.2 0.2

RB 16 July 29-31 6.7 6 44 6.6 6.7 0.492 38 1.9 0.55 0.0059 0.218 -8.4 -16 -7.6 1.3 0.7 0.6

RB 17 August 2-4 7.1 10 54 7.5 7.2 0.491 50 1.5 0.47 0.0062 0.229 -15 -22 -7.8 2.0 1.0 1.1

RB 18 August 5-7 5.7 8 51 7.5 6.8 0.484 65 1.6 0.48 0.0051 0.187 -15 -24 -9.1 2.1 1.0 1.1

RB 19 August 9-11 4.5 9 60 6.6 9.1 0.489 80 1.4 0.54 0.0041 0.150 -17 -23 -6.6 1.9 1.0 0.9

RB 20 August 12-14 3.7 5 53 5.9 9.1 0.484 90 1.6 0.62 0.0035 0.129 -14 -23 -8.4 1.6 1.0 0.6

mean 44 7.4 6.4 0.490S.D. 11 2.9 1.9 0.009

166

Foot notes from Table 3:

a Average water concentrations of 3.2 ng L-1 for a-HCH and 0.35 ng L-1 for g-HCH and EF = 0.444 for a-HCH were averages at stations 7,8,9 and 20 (ref 19); HLCs as in Table S2.

b,c,d See Table S2 for descriptions.

e ND, NV and N refer to deposition, volatilization and net actual fluxes (adjusted for % open water).

Table 4: Water concentrations of HCHs by zonea, fugacity ratios and net fluxesb

Water Air fW/fA Net actual flux (ng m-2 d-1)b

Zonea/Location stationsa stationscmean s.d. mean s.d. mean s.d. αααα-HCH s.d. γγγγ-HCH s.d. αααα-HCH s.d. γγγγ-HCH s.d.

1: east archipelago 1-6,25-30 1-5,23-25 2.4 0.49 0.24 0.03 0.452 0.006 1.5 0.34 0.29 0.12 6.6 4.8 -2.2 2.0

2: central archipelago 7-9,19,20 6-10,18 3.2 0.28 0.36 0.05 0.444 0.003 2.1 0.54 0.40 0.27 4.3 9.3 -0.12 0.10

3: north archipelago 21-24 19-22 4.4 0.32 0.36 0.03 0.441 0.006 3.1 1.1 0.78 0.50 38 22 -1.3 1.7

4: west archipelago 16-18 14-17 4.9 0.44 0.42 0.03 0.436 0.003 2.6 0.39 0.48 0.20 14 20 -1.2 2.1

5: Beaufort Sea 10-15 11-13 4.7 0.58 0.31 0.05 0.436 0.003 2.0 0.14 0.22 0.07 34 12 -7.2 5.1

RB before ice breakup 7-9,19,20 RB 3.2 0.28 0.36 0.05 0.444 0.003 2.4 0.65 0.55 0.23 0 0

RB after ice breakup 7-9,19,20 RB 3.2 0.28 0.36 0.05 0.444 0.003 1.6 0.16 0.55 0.07 6.2 2.5 -0.64 0.35

a: ref 19b: net actual flux: adjusted for fraction of open water.c: paired with water stations to calculate fugacity ratios and fluxes.

γγγγ-HCH, ng L -1 αααα-HCH EFαααα-HCH, ng L -1

167

Table 5: Fugacity and flux calculations for TNW-99

Air/Water Wind Air Open αααα-HCH γγγγ-HCH

Samplesa Speed Temp Water fW/fA b kA

c Actual flux (ng m-2 d-1)e Actual flux (ng m-2 d-1)

(m s-1) oC % αααα-HCH γγγγ-HCH (m s-1) DAW d ND NV N ND NV N

1/3-4 5.6 2.6 100 1.3 0.22 0.0049 0.18 -21 -27 -6.3 5.5 1.2 4.3

2/3-4 8.0 1.6 100 1.1 0.23 0.0071 0.27 -36 -39 -3.3 7.4 1.7 5.7

3/5 7.1 2.3 100 1.4 0.25 0.0063 0.24 -24 -33 -9.1 5.4 1.3 4.1

4/5-6 4.6 5.6 50 1.6 0.21 0.0041 0.15 -7 -11 -4.3 2.0 0.4 1.6

5/6 3.0 6.1 10 1.6 0.19 0.0029 0.11 -1.1 -1.7 -0.6 0.3 0.1 0.2

6/6 1.6 8.9 10 1.5 0.17 0.0020 0.07 -0.8 -1.1 -0.4 0.2 0.0 0.2

7/6 4.4 6.5 10 1.5 0.25 0.0040 0.15 -1.5 -2.3 -0.8 0.3 0.1 0.2

8/6 6.4 6.1 0 2.9 0.33 0.0056 0.21 0 0 0.0 0.0 0.0 0.0

9/7 5.0 7.1 0 2.3 0.33 0.0044 0.16 0 0 0.0 0.0 0.0 0.0

10/8-9 9.2 10.9 5 1.9 0.37 0.0082 0.30 -1 -3 -1.4 0.4 0.1 0.2

11/10-11 6.2 8.9 100 2.1 0.23 0.0055 0.20 -28 -59 -31 7.4 1.7 5.7

12/11-12 4.6 4.5 100 2.0 0.29 0.0041 0.15 -25 -49 -24 4.4 1.3 3.1

13/13 10.4 12.7 100 1.8 0.15 0.0094 0.34 -59 -107 -48 15 2.3 13

14/14-16 15.3 5.8 50 2.2 0.32 0.0147 0.55 -35 -77 -42 6.4 2.1 4.4

15/16 6.5 2.1 0 2.3 0.32 0.0057 0.22 0 0 0.0 0.0 0.0 0.0

16/17 2.8 0.3 0 2.9 0.54 0.0028 0.10 0 0 0.0 0.0 0.0 0.0

17/18-19 5.4 2.3 50 2.9 0.73 0.0048 0.18 -8 -22 -15 1.2 0.8 0.3

18/19 6.1 1.9 100 2.2 0.94 0.0054 0.20 -19 -42 -23 1.9 1.8 0.1

19/20-21 4.6 -0.4 100 3.4 1.42 0.0041 0.16 -11 -36 -25 -1.0 -1.4 -0.4

20/22 4.2 -1.4 100 2.2 0.41 0.0038 0.15 -18 -40 -22 3.2 1.3 1.9

21/22 6.7 -1.1 100 2.3 0.36 0.0059 0.22 -26 -61 -35 5.2 1.9 3.4

22/23 10.1 0.7 100 4.6 0.93 0.0091 0.34 -19 -89 -69 2.8 2.6 0.2

23/25 5.1 0.5 100 1.6 0.41 0.0045 0.17 -17 -27 -10 2.3 0.9 1.4

24/29 5.5 1.6 100 1.5 0.38 0.0049 0.18 -20 -30 -10 2.9 1.1 1.8

25/29 4.5 0.8 100 2.3 0.52 0.0041 0.15 -11 -26 -15 1.5 0.8 0.7

a Air samples paired with water samples from ref 19.

b HLC (Pa m3 mol-1 K-1) is 0.061 for a-HCH and 0.024 for g-HCH, freshwater values increased by 20% for salting-out effect

(ref 13). Water temperature is -1.5oC, the freezing point of sewater at 32 psu.

c kA (m s-1) = 0.001 + 0.000462 x [6.1 + (0.63 x U10)]0.5 x (U10 x 0.4899) (Bidleman, T.F.; McConnell, L.L. A review of field experiments

to determine air-water gas exchange of persistent organic pollutants. Sci. Total Environ. 1995, 159, 101-117.)

d DAW = 86400 x kA / RTA

e ND, NV and N refer to deposition, volatilization and net actual fluxes (adjusted for % open water).

168

Ice cover and net actual fluxes

Fluxes were calculated from air and water concentrations, Henry's law constants, wind speeds and fractional ice

cover using equations from Jantunen and Bidleman (25). The wind speeds were taken on board the ship during

TNW-99 and at the Meteorological Service of Canada weather station at RB.

Although FRs indicate the potential for gas exchange, actual exchange rates in the Arctic depend strongly on

ice cover. Inhibition of gas exchange due to ice was included because the extent of ice varied from open water to

total coverage during the sampling campaign. Daily ice maps for the Canadian Archipelago in summer 1999 were

obtained from the Canadian Ice Service, Environment Canada (http://ice-glaces.ec.gc.ca/). A 2.5o radius (~500 km

diameter) was taken around the sampling point to estimate percent open water (Tables 3 and 5). During the first leg

of TNW-99 in June, the northern Davis Strait around Devon Island, southern Ellesmere Island and the eastern

portion of Lancaster Sound (south and east of RB) were mostly bergy water (<10% sea ice). There was open water

in the Davis and Hudson straits (air samples 1-3) and partial to complete ice cover was encountered from Foxe Basin

to RB (samples 4-10). Leg 2 took place during August where less ice was generally encountered, but complete ice

cover was observed between Banks and Melville Island (samples 15-16) and to a lesser extent during samples 14 and

17 (Tables 3 and 5, and Figure 2).

Figure 2: Sea ice extent in July 1999 and August 1999, showing the break up of ice around Resolute Bay.

169

RB is on the border of the minimum ice coverage which generally occurs in late July to early August. The

ice around Cornwallis Island showed the first signs of breaking up on July 16 and by July 20 the open water in

Lancaster Sound had extended to Barrow Strait. The ice continued to retreat and on July 23 there was open water in

Wellington Channel to the east of Cornwallis Island.

Results are expressed as "net actual fluxes":

))(%( OWNNN vd −= (3)

where Nd and Nv are the deposition and volatilization fluxes (ng m-2 d-1) and %OW is the percent open water. These

estimates are summarized in Table 4, with details in Table 5. Net actual fluxes from water to air are designated as

negative, while those from air to water are positive. During TNW-99, net actual fluxes ranged from -69 to 0.0 ng m-2

d-1 for α-HCH and -0.4 to 13 ng m-2 d-1 for γ-HCH. Net actual fluxes at RB ranged from -9.1 to 0.0 ng m-2 d-1 for α-

HCH and -0.4 to 0.3 ng m-2 d-1 for γ-HCH.

Inhibition of gas exchange by ice cover reduced both the volatilization and deposition fluxes of HCHs to

zero during RB samples 1-12, and exchange of HCHs took place only from the onset of ice breakup through the late

summer (samples 13-20). The FR of α-HCH at RB before ice breakup averaged 2.4 ± 0.64. After ice breakup

volatilization raised the air concentration from 37 ± 9.1 pg m-3 to 53 ± 5.0 pg m-3, a significant (p<0.001) increase of

43% and the FR was reduced to 1.6 ± 0.16. The mean air concentrations of γ-HCH at RB before (8.0 ± 3.8 pg m-3)

and after (6.6 ± 0.76 pg m-3) ice breakup were not significantly different (p = 0.25) nor were the FRs (0.55 ± 0.23

before and 0.52 ± 0.12 after). Thus, the small net actual fluxes of γ-HCH at RB were insufficient to cause noticeable

concentration changes in either the air or water.

Temperature dependence of air concentrations

It is common to use the Clausius-Clapeyron equation (ln partial pressure vs. reciprocal air temperature) to

gauge the relative contributions to air concentrations of diffuse advective sources versus evaporative emissions from

local surfaces (26,27). Making such conclusions from this study was problematic because the TNW-99 ship passed

through waters having varying HCH concentrations (Table 1) and thus varying potentials for exchange. Also, the

exchange was greatly influenced by ice cover. No significant correlation (r2 = 0.06, p = 0.23) was found between

reciprocal air temperature and α-HCH partial pressure during TNW-99 but a significant correlation was found for γ-

HCH (r2 = 0.58, p = 0.003). At RB, a significant correlation was found for α-HCH (r2 = 0.34, p = 0.008), but not for

γ-HCH (r2 = 0.01, p = 0.73), Figure 3.

170

Figure 3: Clausius-Clapeyron plots for αααα- and γγγγ-HCH at RB and TNW-99.

Ratios of αααα-HCH/ γγγγ-HCH

The ratio of α-HCH/ γ-HCH in air is used to follow the transport of technical HCH versus lindane (28,29).

An accepted ratio in the technical mixture is 4-7 (1,30). Lower ratios are often interpreted as indicating co-

transport of lindane. Higher ratios may be due to fractionation of technical HCH during transport, because γ-HCH is

more likely to undergo deposition by precipitation and gas exchange than α-HCH (31). An examination of the July

ratios of α-HCH/ γ-HCH during TNW-99 shows that they span a narrow range from 3.4-4.9 (mean 4.1 ± 0.6) and in

general are lower than most of the ratios in samples collected at RB during the same month (3.4-9.6, mean 6.5 ± 2.1)

(Tables 2-3).

Among the RB samples, lower α-HCH/ γ-HCH ratios and higher concentrations of γ-HCH were observed

in 1,2 and 8-10. Back trajectories (calculated at 925 hPa and 850 hPa every 6 h) at RB showed that the air parcels

for samples 9 and 10 and to a lesser extent 8, originated from the south over Manitoba. The primary use for lindane

in Canada was as a seed treatment on canola, a crop mostly grown in the Canadian prairies, including Manitoba.

Waite et al. (32) calculated forward air trajectories from the prairies for May and June 1997-1998 (canola planting

season) and found that 6% of the time the trajectories reached the high arctic, which includes RB.

Back trajectories for RB samples 1 and 2 had air parcels that tracked back over the Chukchi/Beaufort Seas

and Greenland respectively, which would appear to rule out the prairies transport route for these samples. Although

the ratio in sample 1 was similar to those reported for that region (10), RB sample 2 was taken during a transport

event from the north-east, at a time when the sea around RB was ice covered and volatilization was hindered (see

below).

-24.4

-24.2

-24.0

-23.8

-23.6

-23.4

-23.2

-23.0

-22.8

-22.6

3.50 3.55 3.60 3.65 3.70 3.75

1/T x1000 (K)

Ln P

(( ((-H

CH

/Pa

-22.6

-22.4

-22.2

-22.0

-21.8

-21.6

-21.4

-21.2

3.50 3.55 3.60 3.65 3.70 3.75

1/T x1000 (K)

Ln

P (

-HC

H/P

a

-22.6

-22.4

-22.2

-22.0

-21.8

-21.6

-21.4

-21.2

3.45 3.50 3.55 3.60 3.65 3.70

1/T x1000 (K)

Ln

P (

-HC

H/P

a

-24.0

-23.8

-23.6

-23.4

-23.2

-23.0

-22.8

-22.6

-22.4

3.45 3.50 3.55 3.60 3.65 3.70

1/T x1000 (K)

Ln P

(-H

CH

/Pa

A

C

B

D

-24.4

-24.2

-24.0

-23.8

-23.6

-23.4

-23.2

-23.0

-22.8

-22.6

3.50 3.55 3.60 3.65 3.70 3.75

1/T x1000 (K)

Ln P

(( ((-H

CH

/Pa

-22.6

-22.4

-22.2

-22.0

-21.8

-21.6

-21.4

-21.2

3.50 3.55 3.60 3.65 3.70 3.75

1/T x1000 (K)

Ln

P (

-HC

H/P

a

-22.6

-22.4

-22.2

-22.0

-21.8

-21.6

-21.4

-21.2

3.45 3.50 3.55 3.60 3.65 3.70

1/T x1000 (K)

Ln

P (

-HC

H/P

a

-24.0

-23.8

-23.6

-23.4

-23.2

-23.0

-22.8

-22.6

-22.4

3.45 3.50 3.55 3.60 3.65 3.70

1/T x1000 (K)

Ln P

(-H

CH

/Pa

A

C

B

D

171

The α-HCH/ γ-HCH ratios in air at RB increased abruptly from July 13 onward due to net volatilization of

α-HCH from surface water as the ice broke up. Pre- and post-breakup ratios were 4.2 ± 0.8 (n=3) and 7.8 ± 1.3

(n=10). The α-HCH/ γ-HCH ratios significantly increased throughout the summer; plotted versus Julian day, they

yielded r2 = 0.45 (p<0.001) for TNW-99 and r2 = 0.55 (p<0.001) for RB.

Exchange of αααα-HCH enantiomers

The enantiomers of α-HCH were determined in air and water samples from TNW-99 and air samples from RB.

All water samples showed depletion of the (+) enantiomer, with enantiomer fractions (EFs) that ranged from 0.432-

0.463 and increased from west to east in the Archipelago (Table 2) (19). The α-HCH in air varied from racemic to

the depletion of the (+) enantiomer (Tables 1-2). The EFs of α-HCH in RB samples 1-12 were generally close to or

slightly less than racemic (0.489-0.501, mean 0.496 ± 0.003)(Table 3). A decrease in the EF to 0.468 in air sample

13 coincided with ice breakup and the EF in samples 13-20 averaged 0.482 ± 0.010, significantly lower than the pre-

breakup mean (p = 0.002) (Figure 4). The EF of α-HCH in surface water of the central Archipelago averaged 0.444

(Table 2 and 4) and the breakup of ice allowed this nonracemic α-HCH to volatilize and mix in the air boundary

layer with the nearly racemic α-HCH advected by long-range transport.

At RB, the following equation was used to estimate the fraction (fA) of α-HCH in the air that volatilized

from the water (33):

) - EF(EF

) - EF(EF f

BS

BM= A (4)

where EFM refers to the mixed composition in the overlying air, EFS to the seawater average = 0.444 and EFB to

background air, assumed to be 0.500. Shen et al. (22) found that the EFs of α-HCH from passive samplers located

near large bodies of water (Lake Superior, North Atlantic Ocean) reflected the EF in the water while the α-HCH in

continental samples was close to racemic. The fA calculated from the mean EFM was 32% after ice breakup.

172

Figure 4: EF of α-HCH and concentration of α-HCH at Resolute Bay (pg m-3). Arrow

indicates ice break up.

During TNW-99, the EFs of α-HCH in air were generally farther from racemic than at RB during the same

time period. One reason for this may be the closer proximity of the air sampler to the water (EC and SLU samplers

were ~10 m and ~25 m above the water, respectively). As stated earlier, the EFs of α-HCH in surface water

increased from west to east (Table 2 and 4). On TNW-99, the EFs in air were significantly (r2 = 0.68 and p <0.005,)

correlated to those in water for events in which there was >90% open water, while the EFs in air for events with 0-

50% open water were higher and not correlated with those in water (Figure 5). There is one point (sample 7, Table

2, Figure 5) that seems to be an outlier. This sample was taken in a small open water polynya that was surrounded

by nearly complete ice cover, so the sample was assigned 10% open water. That the EFs of the water and air were

closely matched for this sample implies that a measurable amount of α-HCH was able to volatilize from the small

open water area around the ship. Thus, the "footprint" that is able to affect the composition of air samples collected

over the water needs better definition.

In air over the Great Lakes and Baltic Sea less racemic α-HCH was associated with higher concentrations

of α-HCH (34,35). Ridal et al. (36,37) also observed that nonracemic α-HCH appeared in air over the lake when

FRs predicted volatilization. Bethen et al. (38) observed that EFs of α-HCH in bulk deposition samples taken on the

shores of the North Sea showed greater deviation from racemic during the warmer part of the year (June -

September) than in May or October. Sundqvist et al. (39) predicted that 20-50% of the α-HCH in the boundary layer

originated from the water during the summer months.

0.44

0.45

0.46

0.47

0.48

0.49

0.50

0.51

EF

0

10

20

30

40

50

60

70

pg m

-3

June

7-9

June

10-

12

June

14-

16

July

1-3

June

24-

26

June

21-

23

June

17-

19

July

19-

21

July

13-

15

July

8-1

0

July

5-7

July

15-

17

Aug

2-4

July

29-

31

July

26-

28

July

22-

24

Aug

12-

14

Aug

9-1

0

Aug

5-7

0.44

0.45

0.46

0.47

0.48

0.49

0.50

0.51

EF

0

10

20

30

40

50

60

70

pg m

-3

June

7-9

June

10-

12

June

14-

16

July

1-3

June

24-

26

June

21-

23

June

17-

19

July

19-

21

July

13-

15

July

8-1

0

July

5-7

July

15-

17

Aug

2-4

July

29-

31

July

26-

28

July

22-

24

Aug

12-

14

Aug

9-1

0

Aug

5-7

173

Global warming is predicted to result in extensive loss of ice cover in the Arctic Ocean (40). One

consequence will be an increased role of air-water gas exchange in delivering and removing persistent organic

pollutants (POPs). Enantiomers of α-HCH, and possibly other chiral POPs, can be sensitive and elegant tracers of

this process.

Figure 5: EF in the water versus EF in the air, showing a correlation when >90% open water (r2 = 0.68), but no

correlation when 0-50% open water .

Acknowledgements

This work was supported by the Northern Contaminants Program (NCP) of Indian and Northern Affairs

Canada and the Swedish Natural Science Research Council. We thank the Swedish Polar Secretariat for ship time

on TNW-99, Narwhal Arctic Services for air sample collection at Resolute Bay, and the Polar Continental Shelf

Project for logistical support at Resolute Bay, Nunavut.

174

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138-145. (15) Su, Y.; Hung, H.; Blanchard, P.; Patton, G.W.; Kallenborn, R.; Konoplev, A.; Fellin, P.; Li, H.; Geen, C.; Stern, G.; Rosenberg, B.; Barrie, L.A. Spatial and seasonal variations of hexachlorocyclohexanes (HCHs) and hexachlorobenzene (HCB) in the arctic atmosphere. Environ. Sci. Technol. 2006, 40, 6601-6607. (16) Falconer, R.L.; Bidleman, T.F.; Gregor, D.J. Air-water gas exchange and evidence for metabolism of hexachlorocyclohexanes in Resolute Bay, N.W.T. Sci. Total Environ. 1995, 160/161, 65-74. (17) Hargrave, B.T.; Barrie, LA.; Bidleman, T.F.; Welch, H.E. Seasonality in exchange of organochlorines between arctic air and seawater. Environ. Sci. Technol. 1997, 31, 3258-3266. (18) Weber, J.; Halsall, C.J.; Muir, D.C.G.; Teixeira, C.; Burniston, D.A.; Strachan, W.M.J.; Hung, H.; Mackay, N.; Arnold, A.; Kylin, H. Endosulfan and γ-HCH in the Arctic: An assessment of surface seawater concentrations and air-sea exchange. Environ. Sci. Technol. 2006, 40, 7570-7576. (19) Bidleman, T.F.; Kylin, H.; Jantunen, L.M.; Helm, P.A.; Macdonald, R.W. Hexachlorocyclohexanes (HCHs) in the Canadian Archipelago. 1. Spatial distribution and pathways of α-, β-and γ-HCHs in surface waters. Environ. Sci. Technol. 2007, 41, 2688-2695. (20) Jantunen, L.M.M.; Kylin, H.; Bidleman, T.F. Air-water gas exchange of HCHs in the Southern Ocean and Antarctica. Deep Sea Res. 2004, 51, 2661-2672. (21) Hellström, A.; Nilsson, M.-L.; Svantesson, E.; Kylin, H. Determination of organic contaminants in the biodegradable fraction of source separated household waste. In: Nilsson, M.-L. Occurrence and Fate of Organic Contaminants in Wastes. Thesis 2000, Swedish University of Agricultural Sciences, ISBN 91-576-5759-9. (22) Shen, L.; Wania, F.; Lei, Y.D.; Teixeira, C.; Muir, D.C.G.; Bidleman, T.F. Hexachlorocyclohexanes in the North American atmosphere. Environ. Sci. Technol. 2004, 38, 965-975. (23) Sun, P.; Blanchard, P.; Brice, K.; Hites, R.A. Atmospheric organochlorine concentrations near the Great Lakes: Temporal and spatial trends. Environ. Sci. Technol. 2006, 40, 6587-6593. (24) Cetin, B.; Ozer, S.; Sofuoglu, A.; Odabasi, M. Determination of Henry’s law constant of organochlorine pesticides in deionized and saline water as a function of temperature. Atmos. Environ. 2006, 40, 4538-4546. (25) Jantunen, L.M.; Bidleman, T.F. Air-water gas exchange of toxaphene in Lake Superior. Environ. Toxicol. Chem. 2003, 22, 1229-1237. (26) Hoff, R. M.; Brice, K. A.; Halsall, C. J. Nonlinearity in the slopes of Clausius-Clapeyron plots for SVOCs. Environ. Sci. Technol. 1998, 32, 1793-1798. (27) Wania, F.; Haugen, J.-E.; Lei, Y.D.; Mackay, D. Temperature dependence of atmospheric concentrations of semivolatile organic compounds. Environ. Sci. Technol. 1998, 32, 1013-1021. (28) Hoff, R.M.; Muir, D.C.G.; Grift, N.P. The annual cycle of polychlorinated biphenyls and organochlorine pesticides in air in southern Ontario. Environ. Sci. Technol. 1992, 26, 266-275. (29) Oehme, M. Further evidence for long-range air transport of polychlorinated aromates and pesticides, North America and Eurasia to the Arctic. Ambio 1991, 20, 293-297. (30) Li, Y.F. Global technical hexachlorocyclohexane and its contamination consequences in the environment form 1948 to 1997. Sci. Total Environ. 1999, 232, 121-158.

176

(31) Iwata, H.; Tanabe, S.; Tatsukawa, R. A new view on the divergence of HCH isomer composition in oceanic air. Mar. Pollut. Bull. 1993, 26, 302-305. (32) Waite, D.T.; Hunter, F.G.; Wiens, B.J. Atmospheric transport of lindane (γ-hexachlorocyclohexane) from the Canadian prairies - A possible source for the Canadian Great Lakes, Arctic and Rocky Mountains. Atmos. Environ. 2005, 39, 275-282. (33) Harner, T.; Wiberg, K.; Norstrom, R. Enantiomer fractions are preferred to enantiomer ratios for describing chiral signatures in environmental analysis. Environ. Sci. Technol. 2000, 34, 218-220. (34) Jantunen, L.M.; Helm, P.A.; Ridal, J.J.; Bidleman, T.F. Air-water gas exchange of organochlorine pesticides in the Great Lakes. Atmosp. Environ. 2007, submitted. (35) Wiberg, K.; Brorström-Lundén, E.; Wängberg, I.; Bidleman, T.F.; Haglund, P. Concentrations and fluxes of hexachlorocyclohexanes (HCHs) and chiral composition of α-HCH in environmental samples from the southern Baltic Sea. Environ. Sci. Technol. 2001, 35, 4739-4746. (36) Ridal, J.J.; Kerman, B.R.; Durham, L.; Fox, M.E. Seasonality of air-water fluxes of hexachlorocyclohexanes in Lake Ontario. Environ. Sci. Technol. 1996, 30, 852-858. (37) Ridal, J.J.; Bidleman, T.F.; Kerman, B.; Fox, M.E.; Strachan, W.M.J. Enantiomers of α-hexachlorocyclohexane as tracers of air-water gas exchange in Lake Ontario. Environ. Sci. Technol. 1997, 31, 1940-1945. (38) Bethan, B.; Dannecker, W.; Gerwig, H.; Huhnerfuss, H.; Schulz, M. Seasonal dependence of the chiral composition of α-HCH in coastal deposition at the North Sea. Chemosphere 2001, 44, 591-597. (39) Sundqvist, K.L.; Wingfors, H.; Brorstrom, E.; Wiberg, K. Air-sea exchange of HCHs and PCBs and enantiomers of α-HCH in the Kattegat Sea region. Environ. Pollut. 2004, 128, 73-83. (40) Dalla Valle, A.; Codato, E., Marcomini, A. Climate change influence on POPs distribution and fate: A case study. Chemosphere 2007, 67, 1287-1295. (41) Bidleman, T.F.; McConnell, L.L. A review of field experiments to determine air-water gas exchange of persistent organic pollutants. Sci. Total Environ. 1995, 159, 101-117.

177

Paper 7

Organochlorine Pesticides and Enantiomers of Chiral Pesticides in

Arctic Ocean Water

Liisa M. M. Jantunena and Terry F. Bidleman b

Archives of Environmental Contamination and Toxicology, 1998, 35, 218-228.

a) Department of Chemical Engineering and Applied Chemistry, University of Toronto, 200 College Street, Toronto, Ontario, Canada, M3S 3E5 b) Environment Canada, 4905 Dufferin Street, Downsview, Ontario, M3H 5T4, Canada Contributions: Liisa Jantunen and Terry Bidleman jointly collected, processed and analysed samples on the BERPAC '93 expedition. In 1994 Liisa collected, processed and analysed the samples from the AOS’94 cruise. Liisa wrote the paper with the guidance of Terry. Terry secured funding and provided scientific guidance during every step of the projects.

178

ABSTRACT

In the summers of 1993 and 1994, seawater samples from the surface layer (40-60 m) were collected to determine

the spatial distribution of organochlorine pesticides on expeditions that crossed the Arctic Ocean from the Bering and

Chukchi seas to the North Pole, to a station north of Spitsbergen and then south into the Greenland Sea. Spatial

differences in concentration were found which varied with the pesticide. Heptachlor exo-epoxide (a metabolite of

heptachlor) and α-hexachlorocyclohexane (α-HCH) increased from the Chukchi Sea to the pole, and then decreased

toward Spitsbergen and Greenland Sea. Chlorinated bornanes (toxaphene) followed a similar trend, but levels were also

high near Spitsbergen and in the Greenland Sea. A reverse trend was found for endosulfan, with lower concentrations in

the ice-covered regions. Little variation was seen in chlordane concentrations, although the ratio of trans-/cis-chlordane

decreased at high latitudes. Several of these pesticides are chiral: α-HCH, cis- and trans-chlordane and heptachlor exo-

epoxide. Enantioselective degradation of (-)α-HCH was found in the Bering and Chukchi seas, whereas the (+)

enantiomer was depleted in the Arctic Ocean and Greenland Sea. Enrichment of (+) heptachlor exo-epoxide was found

in all regions. Trans- and cis-chlordane were nearly racemic.

INTRODUCTION

Organochlorine (OC) pesticides are well-known contaminants of the arctic food chain, moving from water into

plankton and bioaccumulated by fish, birds and marine mammals (Hargrave et al. 1992, 1993; Muir et al. 1988, 1992;

Norstrom and Muir 1994). OC pesticides and their stable metabolites are ultimately passed on to aboriginal people

through consumption of traditional foods (Jensen et al. 1997; AMAP 1997). Many OC pesticides and components of

technical pesticide products are chiral and are applied as racemic mixtures in which the ratio of the (+)/(-) enantiomers

(ER) is 1.00. Physical transport processes and abiotic chemical degradation are not enantioselective, but biological

pathways have the ability to alter ER values from racemic. Such pathways include enzymatic degradation (Müller and

Buser, 1994; Faller et al. 1991a; Buser and Müller 1995; Buser and Müller 1993; Hühnerfuss et al. 1992) and

preferential transport through biological membranes (Möller et al., 1994; Hühnerfuss et al. 1993).

Selective degradation or accumulation of single enantiomers may have toxicological implications. For

example, the (-) enantiomer of o,p'-DDT has a higher estrogenic activity than (+)o,p'-DDT (McBlain et al. 1976) and the

(+) enantiomer is depleted in human fat (Müller and Buser 1995). Enantiomers are also useful as marker compounds to

follow transport processes. Microbially processed pesticides often have non-racemic ERs and can be thereby be

distinguished from the freshly applied, racemic compounds.

179

Fig 1: Cruise track of AOS-94. Dots running from the Chukchi Sea to the Greenland Sea correspond to the station numbers in Table 1.

Enantioselective breakdown of α-hexachlorocyclohexane (α-HCH) has been found in Lake Ontario and arctic seas,

where volatilization from water was signaled by the appearance of non-racemic α-HCH in the overlying air (Jantunen

and Bidleman 1996, 1997; Ridal et al. 1997).

In July-September 1994, samples were collected on a transect from the Chukchi Sea to the Greenland Sea across

the North Pole, on the Canadian icebreaker Louis S. St. Laurent (AOS-94). The cruise track is shown in Fig. 1, and

coordinates of stations and hydrographic information are given in Table 1. Measurements of OC pesticides in surface

waters were made to determine their spatial distribution in the western Arctic Ocean. A previous expedition (BERPAC-

93, Jantunen and Bidleman, 1995) covered the Bering and Chukchi seas for HCHs. Air-water gas exchange of HCHs

and the use of α-HCH enantiomers as a tracer of volatilization from arctic waters has been reported previously (Jantunen

and Bidleman 1995, 1996,1997). This article describes the surface water concentrations of HCHs, chlorobornanes

(CHBs, e.g. toxaphene), the cyclodiene compounds trans- and cis-chlordane (TC, CC), trans- and cis-nonachlor (TN,

CN), heptachlor exo-epoxide (HEPX, a metabolite of heptachlor), and endosulfan-I and II (Endo-I and Endo-II). The

ERs of the chiral compounds α-HCH, TC, CC and HEPX are also reported.

EXPERIMENTAL


Water samples of 4-20 L were collected and HCH compounds were adsorbed on solid phase extraction cartridges

180

containing 1-g C8-bonded silica, using procedures described elsewhere (Jantunen and Bidleman 1995). Cartridges were

extracted with ~10 mL of dichloromethane, cleaned up over a 1-g alumina (6% water) column and shaken with 18 M

sulfuric acid. The final volume for analysis was 1.0 mL. Sampling for other OCs was done by collecting ~200 L of

surface water via a submersible pump; the line running from the pump was Teflon surrounded by a metal mesh. Water

was stored in pressurizable stainless steel cans, passed through a glass fiber filter (142 mm Whatman GF/F, nominal

cutoff 0.7 µm) to collect particulate matter and then pulled at 150 mL/min with a peristaltic pump through a 1.5 cm i.d.

glass column containing ~50 mL of precleaned XAD-2 resin to collected the dissolved OCs. After use, the XAD-2 was

transferred to amber bottles with Teflon-lined lids and stored at 4oC.

The filters were cleaned by baking for 12 h at 400oC in a muffle furnace, wrapped in aluminum foil and stored in

plastic bags. The XAD-2 was prepared by sieving under water and precleaned by soxhlet extraction for 24 h each with

the following sequence of solvents: acetone, petroleum ether, dichloromethane, petroleum ether and acetone. The XAD-

2 was stored at 4oC in water prior to preparing the extraction columns.

The XAD-2 resin from water sampling was soxhlet extracted in precleaned thimbles with methanol for 24 h

followed by dichloromethane for 24 h. Both extracts were concentrated to ~150 mL with a rotary evaporator. The

dichloromethane extract and 100 mL water were added to a separatory funnel, mixed well and the dichloromethane was

removed. The aqueous layer was extracted with another 50 mL dichloromethane. The methanol extract was diluted

with 50 mL saturated sodium chloride and 100 mL water and shaken with 50 mL dichloromethane. Extraction of the

water-methanol layer was repeated twice. All dichloromethane extracts were combined and dried over granular

anhydrous sodium sulphate. Water filters were extracted by refluxing for 12 h with 400 mL dichloromethane. The

extracts of the XAD-2 and filters were concentrated separately to 1-2 mL and the solvent was exchanged into hexane by

rotary evaporation and blowing down with a gentle stream of nitrogen. The extracts were cleaned up on a 1-g column of

neutral alumina (deactivated with 6% water,) eluted with 10 mL of 20% dichloromethane in petroleum ether. The eluate

was exchanged into isooctane, cleaned up with 18 M sulfuric acid (omitted for HEPX and endosulfan analysis) and

reduced by nitrogen blowdown to 100-200 µL for analysis. The water and solvents that were used were chromatographic

quality.

18

1

Table 1: Hydrographic Information and Concentrations (pg/L) of Dissolved Pesticides in Surface Water

Station Latitude Temperature Salinity αααα-HCH γγγγ-HCH HEPX TC CC TN CN Endo-I Endo-II Σ Σ Σ ΣCHBs

North Longitude oC (psu) Single RF Multiple RF T2 T12

1 67o47' 168o47' W -0.1 32.69 1700 320 6.6 2.0 1.4 0.88 0.26 5.6 2.8 14 18 0.29 0.36

2 72o08' 168o50' W -1.7 32.14 1290 250 6.0 1.6 0.92 0.63 0.40 8.8 7.8 24 33 0.24 0.37

7 75o 00' 169o 59' W -1.5 30.63 2490 410 na na na na na na na na na na na

11 76o38' 173o19'W -1.6 31.01 2360 370 9.6 1.1 0.94 0.66 0.23 1.2 1.4 26 25 0.40 0.29

13 77o48' 176o18'W -1.6 30.44 2550 360 10.7 1.8 1.3 1.0 0.40 3.1 2.4 37 40 0.64 0.46

16 78o59' 175o49' W -1.6 30.15 2400 400 18.8 1.0 1.0 0.73 0.30 3.1 0.6 34 35 0.81 0.56



20 80o20' 178o38'W -1.6 31.18 2180 400 13.6 0.92 0.91 0.55 0.29 2.3 1.0 50 44 0.77 0.59

24 82o28' 175o40'E -1.7 32.07 2490 470 19.6 0.83 0.88 0.52 0.26 1.9 1.3 81 65 1.0 0.90

25 83o10' 173o56'E -1.7 32.23 2620 580 17.5 1.6 1.5 1.0 0.51 3.6 4.1 74 74 1.1 1.0

26 84o 04' 175o 04' E -1.6 31.05 2070 580 na na na na na na na na na na na

28 85o54' 166o42' E -1.6 31.57 2740 380 12.6 0.70 0.78 0.46 0.26 1.8 1.0 55 46 0.73 0.62

29 87o09' 160o42' E -1.6 32.1 2090 700 17.4 1.4 1.3 0.74 0.33 0.4 0.9 75 52 1.0 0.83

30 88o 04' 174o 50' E -1.7 33.64 2630 520 na na na na na na na na na na na

31 88o47' 142o44' E -1.7 32.58 2430 560 14.2 1.5 1.5 0.91 0.39 1.5 0.6 92 65 1.1 0.97

35 90o00' -1.7 31.89 2690 550 13.7 2.0 2.1 1.4 0.35 1.3 0.1 55 66 0.87 0.41

37 84o15' 35o05' E -1.6 33.41 950 220 8.4 2.3 1.2 1.6 0.49 3.1 0.5 96 96 1.0 0.90

38 83o51' 35o41'E -1.8 34.09 1040 170 na na na na na na na na na na na

39 75o00' 06o03' W 4.0 34.29 630 200 4.7 0.84 0.48 0.42 0.28 4.2 0.2 57 58 0.55 0.45

182

Analysis

Quantitative and enantiomer analyses were done on a Hewlett Packard 5890 GC-5989B MS Engine operated in in

the negative ion mode with methane at a nominal pressure of 1.0 Torr. The ions monitored (target/qualifying) were:

HCHs (255/257), TC and CC (412/410), TN and CN (444/446), HEPX (388/386), OXY (422/420), endosulfan and

mirex (404). The 7-Cl, 8-Cl and 9-Cl homologs of CHBs were determined by monitoring the 343/345, 379/381 and

413/415 ions (Bidleman et al., 1995). The column used for quantitative analysis was a DB-5MS (J&W Scientific,

U.S.A., 30 m x 0.25 mm i.d., 0.25 µm film thickness), operated at a helium carrier gas flow of 40 cm/s. The temperature

program was: initial temperature 90oC, 15oC/min to 160oC, 3.5oC/min to 210oC, hold 1.0 min, 20oC/min to 260oC hold

5.0 min. Sample volumes of 2 µL were injected splitless (split opened after 1.5 min). Other temperature conditions were:

injector and transfer line 250oC, ion source 150oC and quadrupole 100oC. Quantification carried out against five

standards that spanned the concentration range of the samples, using HP MS Chemstation software. Mirex was added to

extracts before injection as an internal standard. Random samples were checked for native mirex and found negative.

The sum of CHBs containing 7-9 chlorines was quantified by two methods based on single and multiple response

factors. The first used the total GC-MS peak area (sum of peaks at all monitored CHB ions, relative to mirex) and the

mass of toxaphene injected. Because the response of CHBs in negative ion MS varies substantially among congeners

(Shoeib et al., 1997), a multiple response factor method was also employed. Response factors were assigned to 41

individual peaks or groups of peaks of the technical toxaphene, similar to the Webb-McCall methods for PCBs (Webb

and McCall, 1973). The mass contribution of peak groups was obtained by GC with flame ionization detection (GC-

FID), which was assumed to respond to the carbon skeleton of the CHBs (Harder et al. 1983). A standard of technical

toxaphene was chromatographed, with hexachlorobenzene and mirex as retention time markers, on a 30 m DB-5 column

using the same temperature program for GC-FID as in the GC-MS analytical method. To correct for injector

discrimination, toxaphene peak areas were normalized to GC-FID response factors (area/ng carbon) for series of PCBs

which spanned the retention time range of toxaphene. The mass contribution of individual peaks was obtained by :

where Ai is the corrected area of peak i, Mi is the molecular weight of compound i (376, 410 or 444 for CHBs having 7,

8 or 9 chlorines) and n is the number of peaks integrated. Negative ion MS response factors (relative to mirex) were

calculated for individual peaks using their percent mass contribution and the total mass of toxaphene standard injected;

84% of the total GC-FID peak area and 60% of the total MS peak area were accounted for. In the case of water

samples, >90% of the total MS peak area was accounted for by the peaks employed in the multiple RF method. Peaks

with retention times matching the two single CHB congeners which bioaccumulate strongly were also determined.

These were 2-exo,3-endo,5-exo,6-endo,8,8,10,10-octachlorobornane, designated as T2 by Stern et al. (1992), P26 by

M A

)(100)M A( = % Mass

ii

n

1=i

ii

∑

183

Frenzen et al. (1994) and B8-1413 by Andrews and Vetter (1995); 2-exo,3-endo,5-exo,6-endo,8,8,9,10,10-

nonachlorobornane also called T12, P50 and B9-1679 by the above authors. Pure standards of T2 and T12 (obtained

from Axact, Inc., Commack, New York, U.S.A.) were used to identify and quantify these two congeners.

Two different chiral columns were used for the enantiomer analysis: Beta-DEX (20% permethylated β-

cyclodextrin in polydimethylsiloxane, 30 m x 0.25 mm i.d., 0.25 µm film thickness, Supelco Corp., U.S.A.) and BGB-

172 (20% tert-butyldimethylsilylated β-cyclodextrin in OV-1701, 30 m x 0.25 mm i.d., 0.25 µm film thickness, BGB

Analytik AG, Switzerland). The Beta-DEX temperature program was: initial temperature 90oC, hold for 1.0 min,

15oC/min to 130oC, 1.0oC/min to 210oC, 20oC/min to 230oC hold for 5.0 min. The BGB-172 temperature program was:

90oC, hold for 1.0 min, 15oC/min to 140oC, 1.0oC/min to 190oC, hold for 2.0 min, 20oC/min to 240oC hold for 5.0 min.

Both columns were operated at a He carrier gas flow of 40 cm/s. Other conditions were: splitless injection (split opened

after 1.5 min), injector temperature 220oC, ion source temperature 150oC, transfer line temperature 220oC and

quadrupole temperature 100oC. Enantiomers of α-HCH and HEPX were separated on the BGB-172 column. The Beta-

DEX column was used as a confirmation column for α-HCH since the elution order of the enantiomers was reversed

from that on BGB-172 (Jantunen and Bidleman 1996, 1997). The Beta-DEX column also resolved the enantiomers of

TC and CC, but Endo-I interfered with the second-eluting (-) enantiomer of CC. A tandem column consisting of Beta-

DEX (30-m) followed by DB-210 (J&W Scientific, U.S.A., 30 m x 0.25 mm i.d., 0.25 µm film thickness) was used to

separate Endo-I from (-)CC. The temperature program for these separations was: initial temperature 90oC, hold for 1.0

min, 15oC/min to 140oC, 0.5oC/min to 210oC, 20oC/min to 230oC, hold for 5.0 min. The tandem column was operated at

a carrier gas flow of 40 cm/s, with other conditions the same. Elution orders of the (+) and (-) enantiomers were

determined by injecting enriched standards of (+) enantiomers, obtained from Axact Inc., Commack, New York, U.S.A.

Quality Control

Detailed information on recovery efficiencies and blanks for HCHs has been reported elsewhere (Jantunen and

Bidleman 1995, 1996, 1997) and is summarized here. Water samples on the AOS-94 and BERPAC-93 cruises were

spiked with HCHs and yielded mean recoveries (n=28) from solid phase extraction cartridges of 71 ± 12% for α-HCH

and 73 ± 14% for γ-HCH. The mean blank values (n=10) for α- and γ-HCH were 0.21 ± 0.03 and 0.12 ± 0.05 ng. From

AOS-94, average blanks (n=3) were: <20 pg HEPX , 26 ± 20 pg CC, 45 ± 24 pg TC, 47 ± 31 pg TN, 70 ± 28 pg Endo-I

and 478 ± 102 pg toxaphene. Two recovery experiments for OCs other than HCHs were done by spiking ~180 L

seawater with 1.8 - 2.5 ng cyclodienes and 100 ng toxaphene, dividing the spike among the stainless steel cans

containing the water, and passing the water through XAD-2. Average recoveries, after adjusting for the native amounts

in seawater, were: chlordanes and HEPX 44%, Endo-I 75% and toxaphene 57%. In previous studies using the same

XAD-2 sampling and extraction procedures, recoveries of TC, CC and TN averaged 56% and toxaphene was 102%

(Bidleman et al. 1995); PCB congeners and p,p’-DDT were 71-94% (Kucklick et al., 1994). Losses were probably

incurred mainly during final evaporation of the samples to 100-200 µL for the GC-MS analysis. This was confirmed by

carrying out laboratory blow-down tests with standards, for which recoveries of cyclodienes and toxaphene averaged

184

50% and 77%. Losses of HCHs were only moderate (recoveries 71-73%, see above) because the final blow-down

volume was 1.0 mL instead of 100-200 µL. Concentrations of OCs in Table 1 have been corrected for recovery factors

of 71-73% for HCHs, 50% for chlordanes and HEPX, and 75% for endosulfans and CHBs.

Quality control issues in enantiomer analysis are precise integration of two peak areas and elimination of

interferences. Racemic standards were repeatedly injected on the two columns to determine the reproducibility of

measuring the area ratio of (+)/(-) enantiomers (ER). Average ER values were: α-HCH = 1.00 and HEPX = 1.01

(BGB-172), α-HCH = 0.99 (Beta-DEX), TC = 0.99 and CC = 0.99 (tandem column), with a relative standard

deviations of ± 0.01 (n = 6-11 for each compound). The criterion used for enantiomer peak purity was agreement of

the target/qualifying ion ratio within ± 5% of the standard values (Falconer et al. 1997). Confirmation of the ERs for α-

HCH in water samples was done by analysis on both the Beta-DEX and BGB-172 columns. The order of enantiomer

elution is reversed on these two column with (+)α-HCH eluting first on Beta-DEX and (-)α-HCH first on BGB-172

(Jantunen and Bidleman 1996, 1997). The average percent difference in ER values between the two columns was 3.6 ±

2.9 % (n=44).

RESULTS AND DISCUSSION

Spatial Distribution of Pesticides in Surface Water

Concentrations of dissolved OC pesticides in surface water on the transect from the Chukchi Sea across the polar

cap to the Greenland Sea are given for each station in Table 1 and summarized as regional averages in Table 2. Trends

with latitude are shown in Fig. 2. The data for HCHs in Fig. 2 include AOS-94 and the Chukchi portion of BERPAC-93

(Jantunen and Bidleman, 1995, 1996, 1997).

In the upper 40 m of the northern Chukchi Sea, α-HCH and γ-HCH averaged 2.06 ± 0.48 ng/L and 0.43 ± 0.09

ng/L (AOS stations 1-2 and BERPAC results, n=21). These compared well with results from a 1988 survey of the same

region (Hinckley et al., 1991; Rice and Shigaev, 1997) and were ~40% higher than concentrations measured at two

Chukchi stations in 1990 (Iwata et al., 1993). In the polar mixed layer (60 m) of the western Arctic Ocean along

AOS-94 track (stations 7-35), α-HCH and γ-HCH averaged 2.42 ± 0.23 ng/L and 0.47 ± 0.11 ng/L. Concentrations

were ~2-3 times lower than these means at stations 37 - 38 near Spitsbergen and station 39 in Greenland Sea, averaging

0.87 ± 0.22 ng/L α-HCH and 0.20 ± 0.03 ng/L γ-HCH. Summarizing, HCHs in the upper 40-60 m increased slightly

along the AOS-94 track from the Chukchi Sea to the pole, then dropped off in the Eurasia Basin and the Greenland Sea

(Fig. 2). Peak concentrations of 2.5 - 2.7 ng/L α-HCH in waters north of 85 are nevertheless lower than those found in

the Beaufort Sea and Canadian Archipelago at latitudes of ~75-81N (3.5 - 7.1 ng/L, Table 2). In recent years water of

Atlantic origin has intruded from the Eurasian side into the northern portion of the Canada Basin (Macdonald, 1996).

The AOS-94 track traversed the region of the Arctic Ocean where the two water types, containing lower and higher α-

HCH concentrations, are mixed.

Spatial trends for other OCs are less well defined due to the limited number of large-volume water samples, and

several of the latitudinal trend bars in Fig. 2 are based on single measurements. HEPX was lower in the Chukchi Sea,

185

near Spitsbergen and in Greenland Sea (6.6 ± 1.9 pg/L, stations 1,2,37,39) but higher in the western Arctic Ocean (14.8

± 3.4, stations 11,13,16,20,24,25,28,29,31,35) (Fig. 2). Measurements of HEPX at the Ice Island (81N, 100W),

Canadian Archipelago, in 1986-87 ranged from <3-11 pg/L (Bidleman et al., 1989; Hargrave et al. 1988) (Table 2).

Chlordanes and nonachlors showed little trend with latitude, other than slightly higher concentrations at stations

35-37 and lower values at station 39. Oxychlordane was detected in some samples at estimated concentrations <0.5

pg/L. The ranges of chlordanes and nonachlors at all stations were (pg/L): TC = 0.70 - 2.3, CC = 0.48 - 2.1 , TN = 0.42 -

1.6 and CN = 0.23 - 0.51. The sum of TC + CC concentrations compared well with the measurements at the Ice Island

during 1986-87 (Hargrave et al., 1988), Resolute Bay in the Canadian Archipelago (74N, 95W) in 1993 (Hargrave et al.,

1997) and the Bering Sea in 1990 (Iwata et al., 1993), all of which ranged from 1.8-3.7 pg/L, but were lower than those

found at Resolute Bay in 1992 (11.8 pg/L, Bidleman et al., 1995) (Table 2). Ratios of TC/CC ranged from 0.9 - 1.9,

with a mean of 1.2 (Table 1). The proportion of TC/CC appeared to be lower in the ice-covered regions, decreasing

from 1.4-1.7 at stations 1-2, to 1.1 + 0.1 at stations 11-35, and then increasing to 1.8-1.9 at stations 37 and 39 (Fig. 2).

Concentrations of Endo-I + Endo-II at stations 1-2 in the Chukchi Sea averaged 12.5 pg/L and were higher than in

the western Arctic Ocean (3.4 + 1.9) or at stations 37 and 39 (3.6 and 4.4 pg/L). By comparison, 2.6 pg/L Endo-I was

reported in surface water from Resolute Bay in 1993 (Hargrave et al., 1997). Endosulfan is used throughout the world

and is one of the few organochlorine insecticides that is still permitted in Canada, the United States and Europe.

Endosulfan in air from the Great Lakes region showed a strong seasonal pattern of high concentrations in summer and

lower in winter (Hoff et al., 1992; Burgoyne and Hites, 1993), a trend also found at arctic air monitoring stations

(Halsall et al., 1998). The proportion of Endo-I to Endo-II in surface water was quite variable, ranging from nearly all

Endo-I (stations 35 and 39), approximately equal concentrations of the two isomers (stations 2, 11 and 25), and more

Endo-II (station 29). Technical endosulfan contains a 2/1 ratio of Endo-I/Endo-II. Endo-I is more volatile and is

usually enriched in air samples (Burgoyne and Hites, 1993). However Endo-II has a lower air/water partition coefficient

(Henry’s law constant, Rice et al., 1997) and therefore should be preferentially deposited by precipitation and air-water

gas exchange. Chan et al. (1994) reported that Endo-II exceeded Endo-I in precipitation samples from the Great Lakes.

Moreover, it has recently been shown that Endo-II can be converted to Endo-I under environmental conditions

(Schmidt et al., 1997). Thus the factors that control the proportion of Endo-I/Endo-II in open-ocean water are complex

and not well understood.

CHBs (Table 1, single response factor values) were lowest in the northern Chukchi Sea (stations 1-2, 14-24 pg/L)

and increased to an average of 58 ± 22 pg/L in the western Arctic Ocean (stations 11,13,16,20,24,25,28,29,31,35).

Stations 38 and 39 were also fairly high (96 and 57 pg/L) CHBs in Resolute Bay were 48 pg/L in August-September

1992 (Bidleman et al. 1995) and followed a seasonal cycle in 1993 from 40-140 pg/L with a yearly average of 85 pg/L

(Hargrave et al. 1997).

186

Fig. 2:Concentration of OCs in the upper 40-60 m of the water column , summarized by latitude (N. HCHs: 65-69 = station 1 + BERPAC-93 data; 70-74 = station 2 + BERPAC-93 data; 75-79 = stations 7,11,13,16; 80-84 = stations 18,19,20,24,25,26; 85-89 = stations 28,29,30,31; 90= station 35; 84-80= stations 37,38; 75 = station 39. Other OCs: 65-69 = station 1; 70-74 = station 2; 75-79 = stations 11,13,16; 80-84 = stations 20,24,25; 85-89 = stations 28,29,31; 90 = station 35; 84-80 = station 37; 75 = station 39. Bar shades are: α-HCH (black) and γ-HCH (white), CHBs: single response factor (black) and multiple response factor (white), HEPX (black), TC (black) and CC (white), Endo-I (black) and Endo-II (white), TN (black) and CN (white) .

65-6

970

-74

75-7

980

-84

85-8

989

-85

84-8

077

-67

0

2

4

6

8

10

65-6

970

-74

75-7

980

-84

85-8

989

-85

84-8

077

-67

0

0.5

1

1.5

2

2.5

3

65-6

970

-74

75-7

980

-84

85-8

989

-85

84-8

077

-67

0

5

10

15

20

25

HEPX and HEPT

pg/L

TC and CC

pg/L

αααα-HCH and γγγγ-HCH

North PoleChukchi Sea

GreenlandSea

ng

/L

65-6

970

-74

75-7

980

-84

85-8

989

-85

84-8

077

-67

0

20

40

60

80

100

120

140

CHBs

pg/L

pg/L

Latitude (N)

pg/L

Latitude (N)

E-I and E-II TN

65-6

970

-74

75-7

980

-84

85-8

989

-85

84-8

077

-67

0

0.5

1

1.5

2

65-6

970

-74

75-7

980

-84

85-8

989

-85

84-8

077

-67

0

0.5

1

1.5

2

2.5

3

187

Table 2: Average Regional Concentrations of Pesticides in Surface Water (pg/L)

αααα-HCH γγγγ-HCH HEPX TC CC TN CN Endo-I Endo-II SCHBs

Bering Sea

1993 2002 454 Jantunen and Bidleman, 1995

1990 1500 190 1.5 1.9 0.50 Iwata et al., 1993.

1988 2400 570 Hinckley et al., 1991.

1988 2440 720 Rice and Shigaev, 1997

Chukchi Sea

1993-94 2060 430 6.3 1.8 1.2 0.76 0.33 7.2 5.3 19 This workand Jantunen and Bidleman (1995)

1990 1400 180 0.90 2.6 0.60 Iwata et al., 1993.

1988 2400 620 Hinckley et al., 1991.

1988 2460 690 Rice and Shigaev, 1997

Western Arctic Ocean

1994 2420 470 14.8 1.3 1.2 0.80 0.33 2.0 1.3 58 This work

1992-1993 2000-3000 500-1100 Macdonald et al., 1996and Jensen et al., 1997

Beaufort Sea and Canadian Archipelago

1993 3640 520 0.5 1.3 2.6 85 Hargrave et al., 1997

1992-1993 3500-5500 500-700 Macdonald et al., 1996and Jensen et al., 1997

1992 4700 450 7.3 4.5 1.5 48 Bidleman et al., 1995

1987 7100 810 Patton et al., 1989

1986-87 4465 610 < 3-11 3.7 (TC+CC) 0.50 145-175a Bidleman et al., 1989and Hargrave et al., 1988

Near Spitsbergen and Greenland Sea

1994 870 200 6.6 1.6 0.84 1.0 0.36 3.7 0.35 77 This work

a) Reported erroneously as 360 pg/L in original paper

The CHB chromatograms were dominated by the more volatile early-eluting congeners, suggesting an atmospheric

source (Fig. 3). The single and multiple response factor approaches were compared for quantification of ΣCHBs. The

ratio of concentrations obtained by the multiple/single methods (Cm/Cs) ranged from 0.69 - 1.38, with an average of

0.99. Regression analysis gave: Cm = 0.73Cs + 10.8, r2 = 0.82. The two methods agreed surprisingly well, considering

that the GC-NIMS response factors for toxaphene congeners vary widely (Shoeib et al. 1997).

Peaks matching retention times of the two highly bioaccumulating individual congeners of toxaphene, T2 and T12,

were also identified in the surface water and were quantified using single-component standards (Table 1). The T12 peak

was the only nine-chlorinated CHB that was detected in substantial amounts. Stern et al. (1992) found that the T2 and

T12 peaks in biological samples were nearly pure compounds, but this is unlikely to be the case in abiotic media.

Shoeib et al. (1997) and deBoer et al. (1997) examined technical toxaphene and found that peaks having the retention

times of T2 and T12 on a DB-5 capillary column could be further separated into several components by multi-

dimensional GC. So the T2 values reported in Table 1 are probably inflated because the peak contains a mixture of co-

188

eluting octachloro- compounds. Likewise, the T12 peak may contain co-eluting nonachloro- compounds. As a percent

of ΣCHBs in seawater, the T2 and T12 peaks averaged 1.5 and 1.3 %. These peaks accounted for 0.52 and 1.2 % of

technical toxaphene (this work, and Shoeib et al., 1997).

The OC pesticides were predominantly found in the dissolved phase, with <0.3-1% on the glass fiber filter for α-

HCHs, <0.5 % for γ-HCH, <0.3-27% for cyclodienes and 2 - 27% for CHBs. Exceptions to this were at stations 16, 20

and 24 , where 16 - 28% of the α-HCH was retained by the filter (but γ-HCH was <0.5%) and station 2 in the Chukchi

Sea, where 32 - 66% of the chlordanes was found on the filter. The latter station was in a region of high productivity

and the water was green with plankton. The stations that showed higher percentages of particulate α-HCH were in a

region of low suspended matter (clear blue water), although under-ice algae were present. The unusually high values for

particulate α-HCH, along with a different degree of enantioselective degradation at two of the three stations (see next

section) suggests that the filters may have collected some plankton, although it is puzzling why the more hydrophobic

pesticides did not also show greater retention on the filter at these locations.

Vertical Profiles of HCHs

In the Chukchi Sea, α-HCH decreased slowly from 2.1 ng/L at the surface to 1.5-1.7 ng/L at 300-350 m, whereas

γ-HCH did not change significantly over this depth range. Concentrations in the 60-115 m layer of the western Arctic

Ocean (2.25 ± 0.24 ng/L α-HCH and 0.52 ± 0.18 ng/L γ-HCH) were not significantly different from average values in

the 60-m polar mixed layer (2.42 ± 0.23 ng/L α-HCH and 0.47 ± 0.11 ng/L γ-HCH) but dropped off rapidly to 0.62 ±

0.24 ng/L α-HCH and 0.26 ± 0.13 ng/L γ-HCH at 200-350 m. The sharp decline of HCHs through the pycnocline is

typical of the Arctic Ocean (Hargrave et al. 1988; Jantunen and Bidleman 1996, 1997; Jensen et al., 1997). At stations

37-38 north of Spitzbergen, α-HCH was 0.95-1.44 ng/L at 10-109 m and decreased to 0.53 ng/L at 235 m and 0.26 ng/L

at 762 m; γ-HCH exhibited no significant trend in the upper 100 m, although a decline was suggested for deeper

samples. HCHs at station 39 in the Greenland Sea were vertically well mixed. Concentrations of α-HCH ranged from

0.59-0.63 ng/L and γ-HCH from 0.17-0.20 ng/L over a depth range of 10-540 m (Jantunen and Bidleman, 1996, 1997).

189

Fig. 3: Chromatograms of the 7-,8- and 9-chlorinated CHBs in surface water at station 37.

Cl-7

Cl-8

Cl-9

Standard

Standard

Standard

Surface Water

Surface Water

Surface Water

24.00 26.00 28.00 30.00 32.00 34.00 36.00

24.00 26.00 28.00 30.00 32.00 34.00 36.00 38.00 40.0042.00

34.00 36.00 38.00 40.00 42.00 44.00

T2

T12

Minutes

190

Enantiomers of Chiral Pesticides

The enantiomer ratios of dissolved α-HCH in arctic waters have been reported by Jantunen and Bidleman

(1996,1997). ERs in surface water were generally >1.00 in the Bering-Chukchi seas, indicating preferential breakdown

of (-)α-HCH.. Depletion of the (+)α-HCH was found in the Arctic Ocean and Greenland Sea, with ERs <1.00. Reasons

for the opposite enantiomer depletion patterns is not known. One hypothesis is that different microbial populations in

these regions are responsible. Faller et al. (1991b) found the (+)α-HCH was preferentially degraded in some regions of

the North Sea whereas (-)α-HCH was depleted in others. ERs were <1.00 in Lake Ontario (Ridal et al. 1997) and a

freshwater arctic lake (Falconer et al. 1995a,b), but >1.00 in soils from British Columbia, Canada (Falconer et al. 1997).

Even though HCHs were >99% in the dissolved phase at most stations (except those discussed earlier) levels of

HCHs in water were high enough to allow ER values to be measured on the filters of the large volume samples. The

particulate α-HCH showed the same, or more, enantioselective degradation than the dissolved fraction (Table 3, Fig. 4).

Both fractions were depleted in the same enantiomer at most stations, but there appeared to be no relationship between

the two and reversal in the depleted enantiomer was found at stations 16 and 29 (Fig. 4). Some particulate samples

showed enhanced degradation of α-HCH compared to the dissolved phase. For example, the ER of dissolved α-HCH at

station 20 was 0.82, whereas the ER on the filter was 0.18 (Table 3). This sample also contained an unusually high

proportion of α-HCH in on the filter, as noted earlier. Station 31 also showed enhanced degradation of particulate α-

HCH (ER = 0.42), but the percent on particles was low (<0.2 %). Other paired dissolved and particulate samples

showed little difference in the ERs. These observations suggest that enantioselective degradation in the dissolved and

particulate phases are decoupled and may involve different microbial populations. The ERs for dissolved α-HCH

generally decreased with depth (Jantunen and Bidleman 1996,1997). At stations 37 and 38 north of Spitsbergen, surface

ERs were 0.82, decreased to 0.64 at 109 m, 0.45 at 235 m and 0.14 at 762 m. Two explanations are suggested to

account for the greater enantioselectivity with depth. As particles settle from the surface, the sorbed α-HCH is

metabolized and released back into the dissolved phase. Alternatively, the ERs may be typical of older Atlantic layer

water which lies below the pycnocline. Distinctive ratios of α-HCH/γ-HCH have been found at different depths in the

North Atlantic near Bermuda, corresponding to water masses of different origin (Fischer and Ballschmiter, 1991).,

however no measurements of ER values have been made in the temperate North Atlantic.

Fig. 5 shows the chromatographic profiles of HEPX, CC and TC enantiomers and the ERs for the dissolved phase

are given in Table 3. Amounts on the filters were too low to obtain good enantiomer signals. HEPX is a metabolite of

heptachlor, so the ER >1.00 probably results from preferential formation of the (+) enantiomer rather than breakdown of

the (-) enantiomer. The range of ERs for HEPX, 1.56-1.76, was tight considering the wide expanse covered by the

stations. Preferential formation of (+)HEPX has been shown to take place on incubation of heptachlor with rat liver

microsomes (Buser et al. 1993) and (+)HEPX enrichment has been reported in salmon, Adelaide penguin, gray seal and

human adipose tissue (Buser et al. 1993; Müller and Buser 1994) and in agricultural soils (Aigner et al. 1998) but the (-)

enantiomer of HEPX predominates in herring and grey seal from the Baltic (Wiberg et al., 1997).

191

Fig. 4: Enantiomeric ratios (ERs) of α-HCH in the dissolved (-) and particulate

(---) phases.

Fig. 5: Chromatograms of HEPX, CC and TC enantiomers in the dissolved phase at station 35.

HEPX

(+)

(-)

TC CC E-I

(+) (-) (+) (-)

1 2 11 13 16 20 24 25 28 29 31 35 37 390.0

0.2

0.4

0.6

0.8

1.0

1.2

1.4

Station Number

ER

of

αα αα-H

CH

192

Table 3: Enantiomeric Ratiosa of Chiral Pesticides in Surface Water

α α α α-HCH α α α α-HCH HEPX TC CC

Station BGB-172 Beta-DEX BGB-172 Tandemb Tandemdissolved particle dissolved particle dissolved dissolved dissolved

Standard 1.00 0.99 1.01 0.99 0.99

1 1.11 1.27 1.08 1.26 1.52 1.00 1.00

2 1.09 1.05 1.09 1.03 1.47 1.00 1.04

11 1.06 nac 1.05 0.98 1.64 1.00 1.06

13 0.95 0.98 0.89 0.93 1.58 1.06 1.02

16 na 1.10 0.83 1.07 na 0.98 na

20 na 0.21 0.82 0.18 na 0.99 na

24 0.92 0.88 0.87 0.95 1.76 0.97 0.99

25 0.96 0.93 0.89 0.95 1.59 0.98 1.01

28 0.92 0.91 0.87 0.89 1.59 1.01 0.94

29 0.90 1.16 0.88 1.23 1.65 0.97 na

31 0.96 0.42 0.90 0.42 1.67 1.01 na

35 0.91 0.84 0.88 0.86 1.63 0.99 1.01

37 0.85 na 0.78 na 1.56 0.97 0.98

39 0.71 0.80 0.68 0.79 1.67 1.03 1.01

a) ER = (+)/(-) enantiomerb) Beta-Dex + DB-210, see textc) na: not analyzed

The origin of HEPX in polar water is uncertain. Photolysis of heptachlor yields racemic photoheptachlor and HEPX

(Buser and Müller 1993). Photoheptachlor has been identified in seal blubber, polar bear fat and human plasma from

the Canadian Arctic (Zhu et al. 1994). However photolysis cannot be responsible for the excess of (+)HEPX found in

surface water unless selective degradation of the (-) enantiomer takes place after formation of racemic HEPX. A more

likely possibility is that atmospheric transport delivers non-racemic HEPX to the Arctic. HEPX with ERs of 1.4-2.0 has

been found in air samples from the southern United States and the Great Lakes regions, apparently the result of

volatilization from soils (Bidleman et al., 1998).

The chlordane compounds TC and CC were close to racemic in the dissolved phase (ER = 0.94-1.06) (Table

3). Enantiomers of both TC and CC are resolved on the Beta-DEX column but Endo-I overlaps (-)CC and the 410/412

ions monitored for chlordanes also give a response for Endo-I. Endo-I was present in sufficient concentrations to make

chiral analysis impossible for (-)CC, so tandem Beta-DEX and DB-210 columns were used to separate Endo-I from (-

)CC as shown in Fig. 5. Tandem columns have been previously employed to separate chlordane enantiomers from each

other and o,p'-DDT enantiomers from p,p'-DDD (Oehme et al. 1994).

Non-racemic residues of α-HCH, HEPX and chlordanes have been found in birds, fish and marine mammals

(Pfaffenberger et al. 1992; Müller et al. 1997; Tanabe et al. 1996, Vetter and Schurig., 1997). These residues may be

193

the result of enantioselective uptake and metabolism or differential permeation of enantiomers through biomembranes

(Müller et al. 1994; Faller et al. 1991a; Buser and Müller 1995; Buser and Müller 1993; Hühnerfuss et al. 1992; Möller

and Hühnerfuss 1993; Hühnerfuss et al. 1994). Reversals in enantioselectivity are sometimes seen between predator and

prey and health status of the animal may be a factor in metabolism of chiral pesticides (Wiberg et al., 1997). The

enantiomeric of composition of OC residues in arctic waters serves as a basis for understanding their uptake and transfer

in the lower food chain.

ACKNOWLEDGEMENTS

We wish to thank crew and scientists of the Russian ship Ocean and the Canadian ship Louis S. St. Laurent for

their assistance in sample collection and the Canadian Department of Indian Affairs and Northern Development,

Northern Contaminants Program, for financial support.

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