Trace Materials in Air, Soil, and Water 9780841231108

Citation preview

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Chapter 1

Measurement of the Trace Element Composition of Airborne Particulate Matter and its Use in Source Apportionment: Case Studies of Lanthanoids and Platinum Group Metals from Houston, Texas Ayşe Bozlaker and Shankararaman Chellam* Zachry Department of Civil Engineering, Texas A&M University, College Station, Texas, 77843-3136 *E-mail: [email protected].

We demonstrate that the composition and abundance sequence for many elements in airborne particles in an urban/industrial environment (i.e. Houston, TX) differs substantially from their average crustal values due to the influence of local anthropogenic sources. Evidence is also provided for the long-range transport of desert dust (i.e. crustal material) and its effects on composition and mass concentrations of ambient particulate matter in Houston using detailed elemental analysis. Lanthanoid metals and platinum group elements (PGEs) are shown to be excellent markers for primary emissions from catalytic cracking operations during petroleum refining and gasoline-driven light-duty vehicles, respectively. Therefore, systematic measurement of numerous representative, transition, and rare earth elements will assist in identifying and apportioning various anthropogenic and natural aerosol sources. Underlying such research are accurate and precise analytical methods to digest airborne particulate matter and quantify trace to major levels of a wide range of elements necessary for robust source apportionment. Our sample preparation procedures and inductively coupled plasma – mass spectrometry techniques for elemental analysis including lanthanoids, rhodium, palladium, and platinum are summarized. We use elemental abundance

© 2015 American Chemical Society In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

profiles and concentration ratios of specific tracer elements, focusing on lanthanoids and PGEs, to track locally emitted and long-range transported aerosols. Finally we list possible sources and their individual contributions to ambient particulate matter (PM) derived from source apportionment studies based on elemental characterization of ambient aerosols at receptor locations.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Introduction Determining the trace element content of inhalable airborne particles is of vital importance because of their potential adverse health impacts (1) and their use in source apportionment (2–4). Metal-bearing particles are emitted by various natural and anthropogenic sources, many of which have characteristic elemental signatures. Air-toxic metals such as As, Cd, Cr, Zn, Pb, etc., are commonly monitored because they are known or suspected to cause serious health effects but can also be used for source characterization (5, 6). However, since these elements are co-emitted by numerous sources, it is difficult to accurately isolate specific human activities or natural sources that release them. For example, Cr is used to make steel and other alloys (7) as well as automobile tire tread rubber (8) and is also present in coal-fired power plant emissions (9). Hence, identifying unique tracers that can pin-point individual sources would assist in accurate source apportionment. Two such examples that our research has focused on are lanthanoids to trace primary emissions from fluidized-bed catalytic cracking (FCC) units of petroleum refineries (3, 10–13). and platinum group elements (PGEs; Rh, Pd, and Pt) emitted from automobiles equipped with three-way catalytic converters (14–16). Lanthanoids are conventionally employed as signatures of geochemical (crustal) processes (17) but can also track primary emissions of fluidized-bed cracking catalysts from petroleum refineries (11–13, 18). Cracking catalysts are strongly enriched in light lanthanoids to impart hydrothermal stability above 750 °C and increase activity, selectivity and gasoline yield (11, 12, 18). In 2010, Texas was home to 27 of the nation’s 137 operating refineries, many of which are located in the Houston Ship Channel, and collectively distilled ~27% of the nation’s crude oil. During refining, ~2,000 tons of catalyst material is estimated to be lost daily by U.S. refineries along with more than 21 tons of lanthanoids through atmospheric emission, incorporation into products, and removal for disposal at landfills. Consequently, aerosols in Houston display a non-crustal lanthanoid signature closely resembling FCC catalysts (10, 19, 20) with smaller contributions from oil combustion and shipping activities (10). These results point to the value of monitoring lanthanoid metals to track primary particulate emissions from refineries. Interestingly, particles emitted from oil-fired plants display similar lanthanoid patterns to those from FCC units since fuel oils themselves contain residues of cracking catalyst materials (18, 21). 4 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Aeolian resuspension of soil dust from the Sahara-Sahel region is estimated to add approximately 800 Tg of dust annually to the atmosphere (22) making it the largest contributor to the world’s dust burden (23, 24). A portion of this mineral material is transported over the Atlantic Ocean to the continental United States (25), increasing ambient airborne particulate matter (PM) levels in Texas occasionally during summer (23, 26). We recently followed a major dust storm originating in Northern Africa that impacted air quality in Houston, TX by more than doubling ambient PM concentrations compared with routine or non-event days. Importantly, such African dust outbreaks dilute anthropogenically emitted elements associated with PM (e.g. Ba, V, Zn, Ni, La) by increasing the mass of crustal components (26, 27). The shift from non-crustal to crustal signatures of lanthanoids in Houston air was used to distinguish particulates that are locally emitted or transported over long distances. A second group of elements, namely Rh, Pd, and Pt (i.e. PGEs) is of interest because of their rarity and high economic value (note that they are classified as “precious metals” along with Ir, Ru, Os, Au and Ag). Rh, Pd, and Pt can be released to the environment from hospitals, industries, municipal wastewater treatment plants, and from high temperature abrasion of autocatalysts (28). PGE signatures depend on their sources, the main one being three-way catalytic converters for gasoline-driven vehicles to oxidize unburnt hydrocarbons and CO and reduce Nox (29, 30). Consequently, it is generally accepted that accumulation of Rh, Pd, and Pt in many urban environments is largely due their release from autocatalysts (particularly in areas where PGEs are not mined or produced) (31). Hence, Rh, Pd, and Pt are being widely used as markers of light duty automobiles in urban environments (32–36). Monitoring these PGEs is especially important to separate mobile source emissions from other stationary sources in heavily trafficked areas such as Texas where over 20 billion vehicle miles were driven in December 2014 alone (37). Measurement of lanthanoids and PGEs along with other main-group and transition elements requires specialized laboratory techniques since they (i) are present only at trace levels and (ii) their complete dissolution from siliceous matrices requires hydrofluoric acid and/or aqua regia (12, 15, 19, 33, 38, 39). These matrices introduce mass spectral overlaps during quadrupole inductively coupled plasma – mass spectrometry (q-ICP-MS) via isobaric and polyatomic interferences, doubly charged species, etc. Therefore extensive sample preparation techniques including 2-stage digestion at high temperature and high pressure as well as matrix separation are necessary. Isobaric interferences and spectral overlaps from polyatomic species arising from the carrier gas (i.e. argon) and the digestion + sample matrix can be reduced by inducing ion-molecule reactions in a collision or reaction cell pressurized with a reactive neutral gas (e.g. NH3) before ions reach the mass spectrometer (40). We have implemented novel protocols based on dynamic reaction cell (DRC) technology in a quadrupole ICP-MS (DRC-q-ICP-MS) to reliably quantify several elements important for air quality investigations e.g. Al, V, Ni, Zn, As, Cu, and Cd (15, 19). PGE quantification requires pre-concentration (e.g. NiS fire assay, Te or Hg co-precipitation) or matrix separation (e.g. ion-exchange) methods before q-ICP-MS (33, 41). To include automobiles in source apportionment calculations, 5 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

ultra-trace levels of PGEs need to be accurately quantified alongside many other elements, necessitating a multistep sample preparation technique, including closed-vessel microwave-assisted acid digestion, repeated evaporation steps, and matrix separation using cation exchange (15). The principal objective of this chapter is to summarize our work on measuring trace element composition of aerosols to identify and apportion their mass concentrations to local and long-range transported sources. Laboratory procedures for the analysis of a wide spectrum of representative, transition, and inner-transition elements by DRC-q-ICP-MS are briefly described. Case studies using lanthanoids and PGEs as unique signatures for tracing emissions of crude oil cracking catalysts, mineral material from the Sahara-Sahel region of North Africa, and gasoline-driven vehicles in Houston, TX are discussed. Finally we identify possible sources of primary particles and calculate their contributions to mass concentrations of ambient aerosols using elemental composition.

Materials and Methods Samples Ambient PM2.5 and PM10 Samplers (Partisol-Plus model 2025, Rupprecht & Patashnick Co.) were deployed at receptor locations in Houston, TX to collect daily airborne PM2.5 (particles ≤ 2.5 µm in aerodynamic diameter) and PM10 (particles ≤ 10 µm in aerodynamic diameter) on 47 mm PTFE membrane filters with a flow rate of 1 m3 h–1. Filters were weighed on three different days pre- and post-sampling using an Orion Cahn C-35 ultra-microbalance (Thermo Electron Corp.) to accurately measure mass of PM collected. More details on sampling locations and sample collection procedures is available in our earlier publications (10, 12, 19, 20, 26, 32).

Vehicular Emissions and Road Dust PM2.5 and PM10 were sampled from inside the Washburn tunnel of Houston, TX (latitude +29.733; longitude –95.211), which runs in the north-south direction underneath the Houston Ship Channel. Approximately 25,000 vehicles traverse the tunnel daily at average speeds in the range 55 – 75 km h–1. The vast majority of these vehicles have only 2 axles and run on gasoline fuel since larger, diesel-driven vehicles are banned as a security measure. PM sampling was performed on a catwalk inside the tunnel, 122 cm above the road surface at a distance of 44 m from the North exit. Three blower fans located on top of the tunnel supply ventilation air, which was also sampled from intake area of the fan room. Paired airborne PM2.5 and PM10 were sampled on 47 mm PTFE filters from inside the tunnel and fan room over 3 – 4 week periods (32, 33). 6 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Road dust material was collected from multiple points in the Washburn tunnel and three major surface roadways in Houston, TX. Samples were dried at 105 °C, sieved through 0.71 mm and 0.106 mm mesh sieves, respectively, and then homogenized as one lot prior to analysis (15, 33).

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Long-Range Transported North African Dust Mineral dust aerosols originating from arid and semi-arid regions of North Africa were sampled at Ragged Point (+13.165, –59.432) located in the East coast of Barbados. Hence, these samples represent aerosols emanating from the Sahara-Sahel region and before they enter the continental United States. Prof. Joseph Prospero from the Rosenstiel School of Marine and Atmospheric Science at the University of Miami has been operating this site for over 4 decades and kindly provided all Barbados samples. Thirteen daily samples for this work were obtained in the months of late May, June, July, August, and mid-September from the years of 2005 – 2008 that coincided with African dust episodes. Total suspended particulates (TSP) samples were collected on Whatman-41 cellulosic filters using a high volume sampler with a flow rate of 1 m3 min–1 (26).

Sample Preparation PGE quantification in environmental matrices require specific sample preparation and analysis techniques of high sensitivity, selectivity and the control of potential isobaric and polyatomic spectral interferences. A multi-step sample preparation procedure, including closed vessel microwave-assisted acid digestion, repeated evaporation steps, and matrix separation by cation-exchange chromatography was employed to extract both PGEs and non-PGEs by modifying a previously reported digestion method used for the extraction of non-PGEs. Details of the procedures are available in our earlier publications (15, 19, 33, 38, 39) and only a brief summary is given below.

Extraction of Non-PGEs Two-stage digestion is necessary for the extraction of a wide range of trace to major elements including lanthanoids. National Institute of Standards and Technology’s (NIST’s) standard reference materials (SRMs), SRM 1648a (urban particulate matter) and SRM 1633b (coal fly ash) were used to develop and validate the method. SRMs were placed in Teflon-lined vessels (HP-500 Plus, CEM Corp.) along with HNO3 and HF, and then digested in a microwave-accelerated reaction system (MARS 5, CEM Corp.) at set points of 200 °C and 300 psig for 20 minutes. HF was used at the 1st stage to completely digest silicate matrix constituents. Next stoichiometric excess H3BO3 was added to each vessel before the 2nd stage to mask any remaining HF and re-dissolve fluoride precipitates, which is particularly important for lanthanoids (39). 7 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

The method was employed on ambient aerosol samples collected on PTFE filters (10, 19, 20, 39). HF and boric acid volumes were proportionately optimized based on the sampled mass. To extract non-PGEs in airborne PM that were collected on Whatman-41 filters, a pre-digestion step was applied before the two-stage procedure to accommodate violent reactions between the concentrated acids and cellulosic substrate of filters. The Whatman filter was first placed in the vessel along with HF and HNO3 and pre-digested in the microwave oven under less aggressive conditions (150 °C, 300 psig, 26 minutes dwell time) (26). Note that only PM samples in Barbados were collected on Whatman filters and not used for PGE measurements. It is emphasized that all blank concentrations were very low compared with sample concentrations allowing facile blank-correction. For example, the ratio of average sample to average blank concentrations for rare earths varied between 31 for Ho, 76 for Dy, 92 for Tb, ~125 for Gd, Pr, and Er, and > 150 for Eu, Tm, Sm, Nd, La, Ce, Yb, and Lu.

Extraction of Both PGEs and Non-PGEs To quantitatively extract all the elements of interest (including Rh, Pd, and Pt), HNO3, HF, HCl, and H3BO3 are all necessary. The formation of nitrosyl chloride by the reactions of HNO3 and HCl is the key step that solubilizes platinum group metals (42, 43).

However, HCl usage coupled with oxygen and nitrogen (from the atmosphere and from HNO3), results in numerous mass spectral interferences, which complicates the measurement of several other marker elements (e.g. Zn, Cr, As, Ba, Cu, Al, Fe, and V). PGE anionic chloro-complexes can be eluted through cationic resins, which retain major interferents allowing sensitive Rh, Pd, and Pt measurements from environmental samples such as road dusts and airborne PM (15, 32, 33, 44, 45). Hence, optimization of the ratio of HNO3 to HCl volumes (aqua regia) is essential for the accurate and simultaneous analysis of platinum group metals and several other elements essential for source apportionment. PGEs from European road dust BCR-723 and spent autocatalyst SRM 2556 were quantitatively recovered without using HF as reported by others (45). Therefore, all metals from ambient aerosol samples were first extracted using aqua regia (200 °C, 300 psig, 20 minute dwell time). Next the digestates were separated into two equal parts for further processing based on analytes of interest. PGE sub-samples were evaporated to dryness, taken into HCl media, and PGEs were then chemically separated from the sample matrix through cation-exchange (Dowex 50WX8 200 – 400 mesh). Sub-samples for non-PGEs followed two-stage acid digestion with HF and H3BO3 (15, 32, 33).

8 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

9

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Table 1. Comparison of Measured Elemental Concentrations with Certified and Indicative Values from European Road Dust BCR-723, Spent Auto-Catalyst SRM 2556, Urban PM SRM 1648a, and Coal Fly Ash SRM 1633b BCR-723 (µg g–1)

SRM 2556 (µg g–1)

Element

Atomic Mass

Supplied value

Measured value

Supplied value

Measured value

Al

27

37,500±2,200

37,307±1,457

400,000

449,418±34,840

Ti

48

2,580±130

2,259±45

V

51

75±1.9

90±2.3

Cr

52

440±18

427±47

Mn

55

1,280±40

1,214±21

Fe

56

32,900±2,000

32,520±921

8,000

8,705±660

Ni

59

171±3.0

167±15

Co

59

30±1.6

30±0.42

Zn

65

1,660±100

1,621±80

Rb

85

75±5

70±1.8

Sr

88

254±19

241±5.5

Y

89

13±1.8

13±0.52

Zr

91

300

317±12

Mo

96

40±0.6

42±2.4

Rh

103

0.0128±0.0013

0.0130±0.0012

51±0.50

55±2.5 Continued on next page.

In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

10

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Table 1. (Continued). Comparison of Measured Elemental Concentrations with Certified and Indicative Values from European Road Dust BCR-723, Spent Auto-Catalyst SRM 2556, Urban PM SRM 1648a, and Coal Fly Ash SRM 1633b BCR-723 (µg g–1)

SRM 2556 (µg g–1)

Element

Atomic Mass

Supplied value

Measured value

Supplied value

Measured value

Pd

106

0.0061±0.0019

0.0052±0.0014

326±1.6

357±22

Cd

112

2.5±0.4

2.6±0.06

Sb

122

28±2.3

26±0.53

Ba

137

460±40

477±14

100

95±13

La

139

7,000

6,894±206

Ce

140

10,000

11,094±296

Hf

178

2.2±0.7

2.0±0.05

Pt

195

0.0813±0.0025

0.0841±0.0058

697±2.3

743±89

Pb

207

866±16

928±61

6,228±49

6,635±152

Th

232

4.8±0.5 SRM 1684a (µg

4.6±0.26 g–1)

SRM 1633b (µg g–1)

Element

Atomic Mass

Supplied value

Measured value

Supplied value

Measured value

Na

23

4,240±60

3,794±272

2,010±30

1,890±175

Mg

24

8,130±120

8,086±277

4,820±20

5,143±452

Al

27

34,300±1,300

35,257±1,810

150,500±2,700

162,183±8,745

Si

28

128,000±4,000

124,531±8,248

230,200±800

259,890±18,043

In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

11

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

SRM 1684a (µg g–1)

SRM 1633b (µg g–1)

Element

Atomic Mass

Supplied value

Measured value

Supplied value

Measured value

K

39

10,560±490

10,770±884

19,500±300

21,283±1,109

Ca

40

58,400±1900

57,222±3,151

15,100±600

16,107±1,587

Ti

48

4,021±86

4,019±209

7,910±140

7,995±692

V

51

127±11

127±7.4

296±3.6

306±17

Cr

52

402±13

407±20

197±4.7

210±12

Mn

55

790±44

799±27

132±1.7

146±10

Fe

56

39,200±2,100

40,059±1,750

77,800±2,300

79,037±3,175

Ni

59

81±6.8

83±4.3

121±1.8

129±11

Co

59

18±0.68

17±0.8

50

60±1.0

Cu

64

610±70

618±42

113±2.6

120±6.8

Zn

65

4,800±270

4,620±457

210

234±21

As

75

116±3.9

127±6.6

136±2.6

145±12

Se

79

28±1.1

26±2.5

10±0.17

11±1.1

Rb

85

51±1.5

50±2.9

140

143±5.1

Sr

88

215±17

224±8.1

1,041±14

1,129±45

Cd

112

74±2.3

74±3.8

0.78±0.06

0.70±0.08

Sb

122

45±1.4

47±1.2

6.0

6.2±0.47 Continued on next page.

In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

12

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Table 1. (Continued). Comparison of Measured Elemental Concentrations with Certified and Indicative Values from European Road Dust BCR-723, Spent Auto-Catalyst SRM 2556, Urban PM SRM 1648a, and Coal Fly Ash SRM 1633b SRM 1684a (µg g–1)

SRM 1633b (µg g–1)

Element

Atomic Mass

Supplied value

Measured value

Supplied value

Measured value

Cs

133

3.4±0.20

3.4±0.40

11

12±0.83

Ba

137

709±27

751±77

La

139

39±3.0

39±1.0

94

94±1.6

Ce

140

55±2.2

56±1.3

190

197±3.8

Sm

150

4.3±0.30

4.3±0.13

20

19±0.64

153

4.1

4.2±0.18

Eu Tb

159

2.6

2.7±0.15

Dy

163

17

17±0.52

Ho

165

3.5

3.0±0.24

Yb

174

7.6

7.5±0.49

Lu

175

1.2

1.1±0.49

Hf

178

5.2

4.6±0.92

6.8

6.0±0.48

W

184

4.6±0.30

4.7±0.53

5.6

5.9±0.66

Pb

207

6,550±330

6,425±255

68±1.1

72±4.4

In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Element

Atomic Mass

Th U

Supplied value

Measured value

SRM 1633b (µg g–1) Supplied value

Measured value

232

26±1.3

25±2.0

238

8.8±0.36

8.9±0.76

Note: Values in bold font correspond to certified values and those in regular font are indicative (uncertified) values.

13

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

SRM 1684a (µg g–1)

In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Analytical Procedure A DRC-q-ICP-MS (ELAN DRC II, PerkinElmer) was used to quantify trace (1 ppb to 100 ppm) levels of PGEs (Rh, Pd, Pt) and lanthanoids (La, Ce, Pr, Nd, Sm, Eu, Gd, Tb, Dy, Ho, Er, Tm, Yb, Lu) along with numerous other trace to major (> 1 %) levels of representative elements (e.g. Li, Be, Na, Mg, Al, Si, K, Ca, Ga, As, Se, Rb, Sr, Sn, Sb, Cs, Ba, Pb), transition metals (e.g. Sc, Ti, V, Cr, Mn, Fe, Co, Ni, Cu, Zn, Y, Zr, Mo, Cd, Hf, W) and actinides (e.g. Th, U) elements. NH3 was used in the DRC unit to overcome spectral overlaps for the isotopes of several elements including Al, Ca, V, Cr, Mn, Fe, Ni, Co, Cu, Zn, As, Se, and Cd. Multielement calibration standards and mixed internal standard of 115In and 209Bi at 20 µg L–1 were prepared before each use by dilution of single-element stock solutions (10 µg mL–1; High-Purity Standards) with 0.4 M HNO3. All calibration blanks and calibration standards for both PGEs and non-PGEs were spiked with the acid matrix containing all reagents employed for digestion at an appropriate amount with the same the dilution factor as the analyzed samples. DRC optimization parameters including the quadrupole dynamic bandpass tuning parameter (RPq) and cell gas flow rate as well as instrumental settings and operating conditions for standard (no NH3) and DRC (using NH3) mode of operations are provided in detail elsewhere (15, 19, 33, 38, 39). Using ammonia as the cell gas suppressed signal intensities by 1 – 2.5 orders of magnitude for Al, V, Cr, Fe, Ni, Cu, and Zn in the blank solution thereby improving the precision and accuracy of their measurements in PM samples (19). More information on improved detection limits and reduced interferences can be found in the original publications (12, 15, 19, 33, 38, 39). Concentrations of principal PGEs interferences (i.e. 63Cu, 65Cu, 68Zn, 85Rb, 88Sr, 89Y, 90Zr, 179Hf, and 206Pb) were reduced by 95.3 to 99.8 % in sample matrices following cation exchange column separation (15). Manufacturer-provided concentrations of all certified and uncertified elements in SRM1648a, SRM 1633b, SRM 2556, and BCR-720 are compared with measured values (from three separate digestions and ICP analysis) in Table 1. Analytical recoveries ranging between 85 – 120 % of supplied values for both PGEs (85 – 110 %) and non-PGEs including lanthanoids (86 – 120 %), demonstrating the accuracy of our laboratory procedures.

Results and Discussion Mass and Elemental Concentrations of Ambient Aerosols in Houston Daily average mass concentrations in the greater Houston area have been reported between 4.7 – 32.1 µg m–3 for PM2.5 and between 12.1 – 125.4 µg m–3 for PM10 (particles ≤ 10 µm in aerodynamic diameter) (10, 19, 20, 26, 46, 47). These fluctuations largely arise from inherently high variability in particulate emissions from myriad natural and anthropogenic sources, proximity of the receptor site to emission sources, and meteorological conditions. Automobile emissions during morning and evening rush hours (32, 48), non-routine emissions 14 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

from local industries (10, 12, 19, 20, 47), regional transport from Mexico and Central America, or trans-Atlantic dust transport from arid/semi-arid regions in North Africa (25, 26) are mainly associated with peak concentrations in Houston. For instance, elevated hourly mass concentrations ~40 µg m–3 for PM2.5 and ~120 µg m–3 for PM10 were measured near the heavily industrialized Houston Ship Channel coinciding with episodic catalyst emissions from crude oil cracking units of local petroleum refineries (10, 19, 20). Similarly, “background” or routine PM2.5 and PM10 concentrations were measured as 13.1 and 36.4 µg m–3, respectively during the non-dust event days in Houston, TX. PM2.5 and PM10 levels more than doubled to 27.8 and 88.7 µg m–3 respectively coinciding with long-range dust transport from the Sahara-Sahel region of North Africa between the dates of July 25 – 27, 2008 (26). Origins of episodic emissions leading to elevated ambient particulate matter concentrations were tracked using concentration ratios, abundance sequences, and enrichment factors of source-specific tracer elements, ternary diagrams, and receptor modeling techniques. Elemental abundance profiles as mass fraction of tracer species including PGEs and lanthanoids along with other “floating” species in representative source material were developed for several primary aerosol sources of particular interest to Houston, TX. These included refinery oil-cracking catalysts (12), trans-Atlantic transported North African dust collected in Barbados (26), tunnel and surface road dusts (15), and vehicular emissions (32, 48). The source profile abundances and the receptor ambient concentrations with appropriate uncertainty estimates were input to the United States Environmental Protection Agency’s (EPA’s) chemical mass balance modeling software (EPA-CMB8.2) to isolate and quantify the contributions of individual pollutant sources to PM2.5 and PM10 mass. The model consists of a system of linear equations that express concentration of each species as a linear combination of source profile abundances and source contributions. Output data includes contribution estimates and their associated standard errors for each source category (49). Previous air quality campaigns in Houston have reported that geochemical and marine markers such as Na, Mg, Al, Si, K, Ca, Ti, and Fe dominates elemental composition of ambient PM2.5 and PM10. Lanthanoids, PGEs and other maingroup and transition metals were also quantified, but at trace (1 ppb to 100 ppm) to minor (0.01 to 1.0 %) levels (10, 12, 15, 19, 20, 26, 32, 47, 50, 51). Among these elements, Ca, V, Cr, Ni, Cu, Zn, Ga, As, Se, Mo, Rh, Pd, Cd, Sn, Sb, Ba, La, W, Pt, and Pb are moderately to anomalously enriched in the proximity of the ship channel indicating “contamination” by local non-crustal sources. The extent of anthropogenic contribution was qualitatively assessed by computing the enrichment factor of individual elements in airborne particles with respect to a crustal reference such as Al, Fe, or Ti (52, 53). A list of non-crustal elements and their corresponding enrichment factors from the Washburn Tunnel and several other receptor locations in Houston are summarized in Table 2.

15 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

16

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Table 2. Enrichment Factors for Non-Crustal Elements in Road Dusts (15) and Airborne Particles (32) from the Washburn Tunnel, Surface Roadways (15), and Ambient Aerosols (10, 26, 32) from Locations near the Houston Ship Channel Washburn Tunnel

Urban/industrial locations

Non-crustal elements

Airborne PM2.5

Airborne PM10

Road dust

Surface roadways dust

Airborne PM2.5

Airborne PM10

Ca

3.4±0.2

7.1±0.3

3.5±0.3

1.0-5.1

0.7-14.4

0.9-19.6

V

0.5±0.3

0.8±0.2

1.4±0.05

0.5-0.9

3.0-86.5

1.8-19.8

Cr

1.9±0.4

1.8±0.4

3.6±0.5

0.5-3.7

0.9-17.3

1.0-15.5

Ni

1.9±1.43

1.2±0.7

1.7±0.2

0.5-0.6

0.8-45.9

1.4-15.1

Cu

227.1±24.1

105.5±15.0

25.3±1.6

0.8-3.4

1.0-132.6

2.1-88.0

Zn

46.5±10.6

65.7±14.6

47.1±3.0

11.5-26.8

5.5-180.1

3.5-100.6

Ga

29.0±4.4

14.5±5.3

17.1±3.0

1.3-8.5

0.6-60.9

0.7-106.2

As

94.5±9.4

89.7±1.6

15.8±0.2

0.8-6.0

5.4-319.1

0.9-187.9

Se

B.D.L.

B.D.L.

B.D.L.

B.D.L.-10.1

50.9-4763

23.2-2239

Zr

9.7±4.5

5.7±1.5

3.8±0.5

1.2-2.3

B.D.L.-6.3

0.7-4.2

Mo

156.1±11.3

51.2±10.4

35.7±5.0

1.2-14.3

10.1-433.9

7.3-178.2

Rh

36575±13500

20996±4722

15700±6236

417.4-884.5

17380-28756

8590-18900

Pd

9142±1708

4250±848.4

2726±876

41.4-427.5

4190-7749

2001-3591

Cd

259.0±116.2

137.1±34.4

46.4±7.8

2.0-13.3

14.9-916.3

8.0-349.2

Sn

381.3±48.6

179.1±31.5

28.6±0.6

7.0-34.8

4.4-726.4

3.2-270.6

Sb

3070±375.3

1342±277.6

211.6±15.1

4.5-17.6

24.9-2820

16.2-1277

In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Urban/industrial locations

Airborne PM2.5

Airborne PM10

Road dust

Surface roadways dust

Airborne PM2.5

Airborne PM10

Ba

14.6±2.2

7.4±3.0

9.8±1.8

0.4-3.8

1.7-30.2

1.2-105.6

La

0.4±0.3

1.0±0.4

1.8±0.3

0.8-1.9

0.5-14.4

0.8-5.7

W

3.5±1.8

5.1±6.1

33.6±5.2

1.3-6.0

1.0-125.8

0.9-108.0

Pt

2549±796.6

889.8±212.3

1642±487.7

125.5-556.0

1853-2376

1119-2107

Pb

10.3±4.2

11.7±4.1

66.6±24.7

2.4-8.0

3.0-62.8

1.6-28.1

B.D.L.: Below method detection limit. Enrichment factors are given either as average ± 1 standard deviation or in a range of minimum and maximum values.

17

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Washburn Tunnel

Non-crustal elements

In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Particulate emissions from single-axle gasoline-driven vehicles in the Washburn Tunnel were extremely enriched in Rh, Pd, and Pt (See Table 2) with respect to their low abundances in the upper continental crust (UCC; 0.018, 0.526, and 0.599 ppb, respectively) (52). Highest enrichment factors were also calculated for PGEs in dusts from surface roadways as well as ambient aerosols just outside the Washburn Tunnel. The results provide a strong evidence for release of Rh, Pd, and Pt from light-duty vehicles and emphasize the importance of their usage as automobile exhaust signatures in urban environments. It is noted that lanthanum was not enriched in the Washburn Tunnel demonstrating that this underwater tunnel is relatively isolated from the Houston atmosphere. However, lanthanum was moderately enriched in ambient PM2.5 and PM10 under the influence of non-routine emissions from nearby petroleum refineries (10, 19, 20). Source apportionment based on detailed elemental composition (including lanthanoids, PGEs, and several others) of Houston-area ambient aerosols is discussed next. Lanthanoids To Track Nonroutine Emissions from Catalytic-Cracking Units Lanthanoid Concentrations in Ambient Particulate Matter The total lanthanoid content (∑14Ln of La to Lu) of ambient aerosols measured in a residential site and three urban/industrial sites in the Houston area has been reported to be between 0.18 – 51.3 ng m–3 for PM2.5 (20, 26, 32) and 0.23 – 20.9 ng m–3 for PM10 (10, 26, 32). These data encompass both routine (i.e. background) and non-routine (i.e. episodic) releases of primary particles from FCC units in local refineries. Episodic air emissions were confirmed with coincident entries in the self-reported emission event database maintained by the Texas Commission on Environmental Quality, ratios of La to other light lanthanoids and La/V, enrichment factors, and abundance sequences. Episodic releases of lanthanoid-enriched catalysts during non-routine operations resulted in 4.4 to 21.4 fold elevation in ∑14Ln levels (7.2 – 51.3 ng m–3) in ambient PM2.5, compared with their corresponding background values (1.6 – 2.4 ng m–3). We also measured higher ∑14Ln concentrations both in ambient PM2.5 and PM10 during an African dust intrusion occurring during a 3-day period (5.0 ± 1.4 and 15.9 ± 3.6 ng m–3, respectively) compared to five-day averages immediately prior to and after the peak (0.85 ± 0.42 and 3.4 ± 2.2 ng m–3, respectively) (26). Lanthanoid-bearing aerosols associated with high mineral material content were distinguished from their anthropogenically emitted counterparts by their rare earth distribution patterns, which were nearly identical to the lithosphere. It was also revealed that ∑14Ln content of PM10 during dust intrusions was controlled primarily by the amount of mineral material.

Lanthanoid Pattern Cracking catalysts have a significantly different lanthanoid signature compared with crustal material, which enriches La but depletes Ce and other 18 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

heavier lanthanoids during episodic releases from petroleum refineries in Houston (10, 19, 20). Similar trends have also been reported by others following transient refinery pollution events (13, 54, 55). Particulate emissions from oil-combustion activities often display similar lanthanoid patterns to those of cracking catalysts since fuel oils contain residual amounts of refining catalysts (18, 21). However, contribution of La-rich particles from oil combustion was found to be at a smaller scale since lanthanoids were strongly and positively correlated with catalytic-cracking operations and crustal material, restraining their releases from oil-combustion activities (10, 12). Additionally, these two sources that are relatively identical in their lanthanoid signatures can be separated using La/V ratios (10, 13, 56).

Figure 1. (a) Lanthanoid abundances in zeolite-based fluid cracking catalysts (12), local soil (15) and UCC (53). (b) Lanthanoid abundances in FCC and local soil that were normalized to those reported values for UCC (53). Error bars represent one standard deviation of the average. 19 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

20

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Table 3. Tracer Elemental Ratios in Ambient PM Collected during Episodic Releases from Refinery Units, Intrusions of North African Dust, and Non-Event (i.e. Routine or Background) Days in Houston (10, 20, 26). Average UCC (53), Cracking Catalysts (12), Refinery Stack and Oil-Fired Power Plant Emissions (3), Vehicular Emissions (32), and Road Dust (15) Are Also Provided for Comparison Ambient PM2.5

Ambient PM10

Ambient PM2.5

Ambient PM10

Ratios

Refinery event

Non-event

Refinery event

Non-event

African dust event

Non-event

African dust event

Non-event

Local road dust

La/Ce

3.2-5.0

0.67-0.89

0.98-9.9

0.42-0.92

0.51-0.74

0.67-6.8

0.49-0.66

0.56-2.8

0.42±0.08

La/Pr

14.3-20.0

6.3-9.3

4.3-58.7

4.0-12.2

4.5-7.0

6.3-40.2

4.4-6.0

5.1-26.9

6.2±0.75

La/Nd

7.6-11.8

1.6-2.6

2.0-21.2

1.0-3.1

1.2-1.9

1.6-22.7

1.2-1.6

1.4-8.0

1.6±0.22

La/Sm

41.1-70.6

8.0-14.3

9.7-108.2

4.0-18.8

6.5-9.6

8.0-60.3

6.3-8.7

7.1-39.3

7.3±1.2

La/V

0.22-1.8

0.02-0.07

0.01-1.1

0.01-1.1

0.06-0.23

0.01-0.25

0.14-0.25

0.07-0.32

0.64±0.12

UCC

FCC emissions

Oil combustion

Vehicle emissions

Ratios

FCC

PM2.5

PM2.5-10

PM2.5

PM2.5-10

PM2.5

PM10

African dust in Barbados

La/Ce

4.3±4.6

1.2

1.3

1.4±0.15

2.5±1.1

0.62±0.25

0.44±0.04

0.49±0.01

0.5

La/Pr

9.7±5.2

N.R.

N.R.

N.R.

N.R.

12.5±3.9

6.9±0.38

4.2±0.07

4.4

La/Nd

6.4±4.1

1.8

1.9

3.3±1.9

3.2±1.4

4.0±0.82

2.0±0.04

1.1±0.02

1.1

La/Sm

55.2±23.3

19.4

19.5

28.5±9.3

30.6±13.0

13.0±1.0

8.7±0.66

5.9±0.16

6.6

La/V

131.3±69.6

13.2

17.6

0.02±0.01

0.25±0.18

0.24±0.06

0.33±0.03

0.19±0.08

0.32

N.R. Not reported. The ratios are given either as average ± 1 standard deviation or in a range of minimum and maximum values.

In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

We also determined the elemental composition of source samples (five fresh and one spent or equilibrium zeolite-based FCC catalysts). Spent catalysts are those that are deactivated following their repeated use in the FCC unit but cannot be further regenerated due to extensive poisoning possibly due to coke, vanadium, and nickel accumulation. Ultimately, the spent catalysts have to be disposed externally, after being classified as a “hazardous waste” possibly in an appropriate landfill. These results along with a FCC source profile as mass fraction of each species can be found in our earlier studies (12, 26, 38). Lanthanoid abundance patterns in catalysts (12) and in local soil (15) before and after normalization to average UCC values (53) are depicted in Figure 1a and 1b, respectively. It can be seen in Figure 1a that crustal material and local soil follow the Oddo-Harkins rule whereas catalytic-cracking catalysts exhibit a different trend (e.g. 57La > 58Ce in catalyst but 57La < 58Ce in the UCC). Therefore, the distinct positive La anomaly in cracking catalysts relative to its crustal abundance makes it a strong tracer for its anthropogenic origins in locations influenced by refinery emissions. We measured light lanthanoids i.e. La, Ce, Pr, Nd, Sm, and Gd, abundances in cracking catalysts to be 30±14 (Sm) to 267±90 (La) times greater than those in the Earth’s crust. Similarly Eu and heavy lanthanoids, i.e. Tb, Dy, Ho, Er, and Yb were also enriched but to a lesser degree (< 12-fold). In other words, cracking catalysts are highly enriched in light lanthanoids and to a lower extent in heavy lanthanoids. Importantly, cracking catalysts are depleted in Ce, Eu, and odd-numbered heavy lanthanoids (Tb, Ho, Tm and Lu) depicting a distinct negative anomaly. As shown in Figure 1a, the lanthanoid pattern in local soil was similar to the average UCC. Hence, the UCC normalized lanthanoid profile of local soil was relatively flat, only exhibiting very weak excursions of La, Gd, Dy, Er and Yb (Figure 1b). Hence, measurement of the complete rare earth signature and Coryell-Masuda diagrams are useful tools to separate natural and anthropogenic emissions of lanthanoid-bearing PM.

Using Ambient Data for Source Characterization Concentration ratios of La to other light lanthanoids (i.e. Ce, Pr, Nd, and Sm) and V in ambient PM2.5 and PM10 during air emission events from petroleum refineries and Saharan dust intrusions are summarized in Table 3. Corresponding ratios for non-event or routine days and several lanthanoid-bearing aerosols from natural and anthropogenic sources are also shown for the sake of comparison. Since both refinery and oil combustion emissions have elevated ratios of La to light lanthanoids, the La/V ratio was used to distinguish the samples affected by these two sources (11, 18, 55, 56). Particulate emissions from fuel and petcoke combustion exhibit low La/V ratio (< 0.1) (55) with respect to UCC (0.31) (53) since they are rich in vanadium. The La/V ratio reaches as high as 13 in refinery stack emissions (3) and becomes an astonishingly high 131.3 ± 69.6 in cracking catalysts (12) due to their highly elevated La content. Also as given in Table 3, La to light lanthanoid ratios for the UCC closely overlap with those measured for mineral dust aerosols that originated in arid regions of North Africa and transported 21 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

across the Atlantic Ocean to Barbados and for local road dust (La/Ce = 0.42 – 0.50). Vehicular emissions display Ce enrichment vis-à-vis La similar to crustal materials, but can exhibit slightly higher ratios (La/Ce = 0.44 – 0.69).

Figure 2. La-Ce-V (a) and La-Ce-Sm (b) ternary diagrams for ambient PM2.5 and PM10 (10, 20, 26) under the influence of routine and non-routine emissions of refinery catalytic cracking units (3, 12, 57) and oil combustion activities (3, 58) in Houston. Other natural and anthropogenic sources (15, 26) of lanthanoid and vanadium bearing particles are included for comparison. Values were adjusted so that the UCC (53) appears in the geometric center of the figure. 22 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

As shown in Table 3 aerosols collected during episodic events were isolated from those collected during routine operations of refineries based on their La-enrichment; i.e. elevated La to other light lanthanoid ratios relative to the UCC and local soil. Similarly, higher La/V ratios in PM2.5 were measured during the refinery emission events (0.22 – 1.8) compared with those collected when no episodic events were self-reported (0.02 – 0.07). Note that during peak event days, atmospheric La/V is significantly higher than the crustal value (0.31) and during routine operation days, atmospheric La/V is significant lower than the crustal value. Therefore, fine PM is predominantly influenced by anthropogenic sources, with strong La enrichment from episodic releases of cracking catalyst. On the other hand, during routine FCC unit operation, V is enriched in the atmosphere due to extensive oil combustion emissions in the heavily industrialized ship channel area, significantly decreasing La/V compared to its crustal value. In contrast, the La/V ratio for PM10 remained in the same range (0.01 – 1.1) for both event and non-event samples. This indicates that in addition to refineries and oil combustion, resuspended crustal material contributed to the lanthanoid atmospheric chemistry preferentially for the coarse size mode in Houston. La to other light lanthanoid ratios in ambient PM collected during an African dust outbreak in Houston, TX were similar to the UCC and African dust aerosols collected in Barbados (e.g. La/Ce < 1, 6.3 < La/Sm < 9.6). As expected, lanthanoid atmospheric chemistry reflected a combination of crustal and refinery sources on days where African dust did not impact Houston (e.g. 0.56 < La/Ce < 6.8, 7.1 < La/Sm < 60.3). It is emphasized that peak-days associated with North African dust increased the total aerosol mass concentration and consequently the total lanthanoid content (∑14Ln). However, it substantially reduced La/Ce (and other La to light lanthanoids ratios) demonstrating a significant change in lanthanoid composition in Houston’s atmosphere, which is otherwise dominated by refinery emissions. Influence of lanthanoid and vanadium bearing aerosol sources on ambient PM2.5 and PM10 (10, 20, 26) is further demonstrated in Figure 2 in the form of ternary diagrams. Three component La-Ce-Sm (Figure 2b) and La-Ce-V (Figure 2a) diagrams were used to distinguish oil combustion and refinery emissions to ambient aerosols. Elemental concentrations were normalized to place average UCC abundances (53) of three components at the centroid. In both diagrams, mineral materials such as North African dust collected in Barbados (26) and local soil (15) grouped around the centroid reflecting their similar La, Ce, Sm, and V composition. Elemental composition of vehicular PM2.5 and PM10 and road dust collected from inside the Washburn Tunnel (15, 32) showed similarity to the UCC with Ce enrichment relative to La. Refinery cracking catalyst (12), FCC unit stack emissions (3, 57), and oil combustion fly ash (3, 59) crowded together near the La-apex in Figure 2b, corresponding to their La-enrichment. In such a representation, PM associated with oil combustion and shipping activities (3, 59) in Figure 2a clustered around the V-apex demonstrating their strong V-enrichment. Elemental composition of ambient PM2.5 and PM10 under the influence of refinery FCC unit emission events and North African dust outbreaks in Houston (10, 26, 39) were included to identify variations in elemental composition under different source impacts. In the La-Ce-V diagram (Figure 2a), 23 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

data for airborne particulate samples enriched in V moved away from UCC, local soil, African dust aerosols, and cracking catalysts toward the V-apex, clustering around the V-rich sources of oil combustion and shipping emissions. Relative La-Ce-V data for some particulate matter samples moved between the V-apex and other sources such as local soil, African dust, and vehicular emissions, reflecting their lanthanoid and V contributions. In the La-Ce-Sm diagram (Figure 2b), these three light lanthanoid concentrations for some ambient particulate matter samples clustered around the centroid demonstrating their crustal origins. However, several others moved away from the centroid toward the La-apex indicating their origins in petroleum refining and oil combustion.

Platinum Group Elements for Tracing Light-Duty Vehicular Emissions PGEs Levels in Road/Tunnel Dust and Airborne Particulate Matter Concentration ratios and levels of Rh, Pd, and Pt in tunnel dust, tunnel gutter, surface roadways dust (15), and airborne PM2.5 and PM10 from inside the Washburn Tunnel of Houston, TX and those in ventilation (ambient) air (32) are summarized in Table 4. Elemental compositions of each samples matrices and source profile abundances as percent mass fraction of all elements quantified in road dusts, airborne PM2.5, PM10, and PM2.5-10 emissions from the Washburn Tunnel can be provided in detail from our previous studies (15, 26, 32) PGE composition of average UCC (52) and a composite recycled catalyst from multiple manufacturers and model types obtained from Engelhard Corp. (currently BASF) (15) in the U.S. are also included in Table 4. Average abundances of Rh, Pd, and Pt dust in the tunnel catwalk ranged between 0.12 – 0.21 µg g–1, 0.59 – 1.0 µg g–1, and 0.41 – 0.67 µg g–1, respectively. Particles collected from the gutter had higher Rh and Pt abundances (0.27 and 1.1 µg g–1, respectively), possibly due to their accumulation via regular washing of tunnel walls and catwalk. In contrast, lower abundance of Pd in the gutter is attributed to its higher water solubility and mobility (60). PGE abundances displayed relatively high variation in dusts collected alongside three major surface roadways. Rh, Pd, and Pt abundances in these samples were lower than those from inside the tunnel, ranging between 0.006 – 0.008 µg g–1, 0.03 – 0.09 µg g–1, and 0.09 – 0.13 µg g–1, respectively. Lower PGE abundances in surface roadways compared with the tunnel was attributed to dilution in the ambient environment due to meteorology, rainfall, mass contribution from other non-PGE sources as well differences in vehicle fleet and driving habits. Average Rh, Pd, and Pt concentrations in ambient PM2.5 and PM10 in tunnel ventilation air were 1.5, 11.1 and 4.5 pg m–3 and 3.8, 23.1, and 15.1 pg m–3, respectively. Similar to concentration differences between tunnel and surface roadways, Rh, Pd, and Pt were significantly elevated in airborne particles inside the Washburn Tunnel compared with ventilation air. Average Rh, Pd, and Pt concentrations increased 4.0- to 9.6-fold inside the tunnel, reaching 12.5, 91.1, and 30.1 pg m–3 in PM2.5 and 36.3, 214, and 61.1 pg m–3 in PM10, respectively. 24 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Table 4. PGE Levels and Tracer Element Ratios in Different Sample Matrices Abundance (ng g–1)

25

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Sample type

Ratios

Dust (15)

Rh

Pd

Pt

Pt/Rh

Pd/Rh

Pt/Pd

Washburn Tunnel dust

152±52.3

770±208

529±130

3.6±0.43

5.2±0.50

0.69±0.02

Washburn Tunnel gutter

273

717

1079

4.0

2.6

1.5

Surface road dust

6.8±1.3

53.5±30.2

106±21.5

15.9±4.7

8.2±5.3

2.5±0.85

Concentration (pg m–3)

Airborne PM (32)

Ratios

Washburn Tunnel PM2.5

12.5±5.8

91.1±28.6

30.1±11.5

2.4±0.21

7.5±1.2

0.33±0.02

Washburn Tunnel PM10

36.3±21.4

214±119

61.1±32.3

1.7±0.12

6.0±0.25

0.29±0.01

Ventilation air PM2.5

1.5±0.5

11.1±4.4

4.5±0.7

3.1±0.56

7.5±0.59

0.43±0.11

Ventilation airPM10

3.8±1.2

23.1±4.1

15.1±3.2

4.0±0.44

6.2±0.89

0.65±0.02

Abundance (µg g–1)

Source

Ratios

Autocatalyst (Engelhard)

184

1148

814

4.4

6.2

0.71

UCC (52)

0.000018

0.000526

0.000599

33.3

29.2

1.14

In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

PGEs abundances in both tunnel and ambient environments were 63 (Pd in surface roadways dust) to 18,760 (Rh in tunnel airborne PM2.5) times higher than UCC levels. Enrichment factors were also calculated for all elements quantified in road/tunnel dust, tunnel aerosols, and aerosols present in ventilation air using Ti as a crustal reference (see Table 2). The PGEs exhibited highest enrichment in all the sample matrices from the Washburn Tunnel (with enrichment factors ranging from 890 for Pt in PM10 to 36,575 for Rh in PM2.5), providing strong evidence for their release from gasoline-driven light duty vehicles.

Figure 3. Three component diagram demonstrating the separation of United States autocatalyst (15), road dust (15) and airborne particulate matter from inside and outside the Washburn Tunnel (32) from European autocatalyst (ERM-EB504), Austrian road dust (BCR-720), and the UCC (52). Houston surface roadways dusts grouped around Canadian autocatalyst, SRM-2556 (used autocatalyst). PGEs contents in various road dusts, roadside soils, and ambient particulate matter reported around the world were taken from the literature (14, 34, 61–71). Values were adjusted so that the average UCC composition52 appears at the centroid. 26 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Rh, Pd, and Pt Composition in Road Dusts and Airborne Particles Influence of light-duty vehicles’ emissions on PGE composition of airborne particles is further illustrated in Figure 3, which simultaneously captures Rh, Pd, and Pt composition in different environmental matrices from around the globe. As before, concentrations were normalized to average UCC values so that the crustal PGE composition appears at the centroid. The United States recycled mixed-lot autocatalyst, Houston tunnel dust (15) and PM2.5 and PM10 emissions from inside the Washburn Tunnel (32) were enriched in Rh with respect to average UCC as evidenced by their position near the Rh-apex. The clustering of these three samples is highlighted by encompassing them in a blue circle. Also as given in Table 4, the Pt/Rh, Pd/Rh, and Pt/Pd concentration ratios for tunnel particles overlapped with those for the U.S. autocatalyst. This provides strong evidence that the primary PGE source is emissions associated with wear and tear of the light duty vehicles’ catalytic converters. Relative Rh, Pd, and Pt concentrations of ambient PM2.5 and PM10 from Houston, TX, Boston, MA and Mexico (14, 32, 61) also grouped in the vicinity of the U.S. autocatalyst, indicating their similar origins. Road dusts from major surface roadways in Houston displayed similar PGE patterns to those from other road dust and roadside soil samples from the U.S. (34, 62, 63), gathering around an older used autocatalyst material, SRM-2556. These samples are grouped within the brown oval. Their common feature is that they were all depleted in Rh but enriched in Pt with respect to freshly emitted autocatalyst material, probably as result of weathering processes in the ambient environment (60). Interestingly, several European samples grouped together and were separated from American samples and are depicted with a red oval. The European autocatalyst (ERM-EB504) was enriched in Rh and Pt, but depleted in Pd with respect to the U.S. autocatalyst. It clustered along with European road dust (BCR-720) and ambient PM from Sweden (64, 68), suggesting their common origin. Ambient PM from Spain (65), Germany (69, 71), and Italy (64, 67), were spatially separated from the European autocatalyst and road dust demonstrating differences in their PGE composition. These variances in PGE composition of road dusts, roadside soils, and airborne particles around the world can be attributed to geographical and temporal differences in autocatalyst composition, regional differences in vehicle fleets and other driving habits, climactic differences, and changes in their environmental transport and fate. Qualitative information on particulate matter sources identified using elemental ratios, ternary plots, and abundance sequences was further probed quantitatively using source apportionment tools as discussed below.

27 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

28

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Table 5. Summary of Houston Area Source Apportionment Studies Using Aerosol Elemental Composition Range of source contribution estimates (%) Source type

PM2.5

PM10

Source signatures

North African dust, non-dust event (26)

1.3-25.4

2.0-31.5

Si, Ai, Fe, Na, Ca, Mg, K, Ti

North African dust, dust event (26)

35.3-67.3

36.7-76.8

Soil and road dust (10, 12, 20, 26, 47, 50)

2.4-51.0

9.9-39.8

Si, Mn, Ca, Fe, Ti, Mg, K, Al, Cs, Ba, Pb, Zn, Ni, Cr, V, Cu, La, Sm, Dy, Er

Si-rich source (10)

N.A.

24.5

Si, Zr, Sr, Cr

Ca-rich source (26, 47)

2.4-33.5

2.6-41.4

Ca

Motor vehicle (20, 26, 47)

2.8-35.7

0.8-30.0

Cr, Mn, Fe, Cu, Zn, Rh, Pd, Pt

Petroleum refineries, non-event (10, 12, 26)

0.6-2.1

0.3-2.3

La, Ce, Pr, Nd, Sm, Gd, Yb, Al, Si

Petroleum refineries, FCC event (20)

12.0

N.A.

Oil combustion & shipping activities (10, 12, 20, 26, 47)

0.4-11.0

0.2-35.0

V, Ni, Co, Sc, Mo

Coal combustion & high temperature operations (10, 12, 20, 26)

1.6-13.5

0.6-9.6

Se, As, Cd, Sn, Sb, Pb, Zn, Si

Industrial combustion (50)

16.9-17.6

N.A.

Zn, Cu, Ni, Fe, Si, K, Ca

Sea salt (12, 20, 26, 47, 50)

0.4-19.0

0.3-14.8

Na, Mg

Vegetative burning (20, 26, 47, 50)

0.9-18.6

N.D.

K

N.A.: Not analyzed, N.D. Not detected.

In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

Apportionment of Primary Particulate Mass to Emission Sources Several investigators have used receptor modeling techniques based on elemental composition to determine that motor vehicles, petroleum refineries, oil combustion and shipping activities, coal combustion and other high temperature operations, vegetative burning, marine aerosols, resuspension of crustal material and road dust, and long-range transported North African dust all contribute to ambient PM2.5 and PM10 mass in Houston to different degrees (10, 12, 26, 32, 39, 47, 50). The number of sources, source tracers, and source contribution estimates are given in Table 5. Local and long-range transported mineral dust sources including North African dust, resuspended soil and road dust, and Ca-rich material are important contributors to both fine and coarse particles (26) During a 3-day outbreak in the year 2008, trans-Atlantic transport of North African dust accounted for more than half the PM2.5 and PM10 mass at two sites in the vicinity of the ship channel (i.e. Clinton Drive and Channelview). During this 3-day peak dust period, crustal elements such as Si, Al, Fe, Na, Ca, Mg, K, and Ti dominated Houston’s atmosphere, significantly reducing enrichment of all anthropogenic metals including lanthanoids referenced to the UCC. Importantly, evidence was provided from chemical mass balancing that small amounts of African dust may remain for several days before and after the peak event. Numerical modeling using a wide suite of major to trace elements also identified a calcium-rich material, potentially emanating from cement plants and gypsum material used to patch parking lots as major aerosol contributors (26, 47). Mineral materials such as resuspended combined soil and road dust (with Mn, Ca, Fe, Ti, Mg, K, Al as signatures), road dust (with Ba, Pb, Zn, Ni, Cr, V, Cu, La, Cs, Dy, Er) preferentially strongly contributed to PM10 (10, 12, 26, 47, 50). A Si-rich material that was co-emitted with Zr, Sr, and Cr was isolated for PM10. Emissions from cat-cracking units can also elevate Al and Si levels (20) in PM2.5 due to the aluminosilicate (zeolite) base of catalysts. Pure siliceous material, not measured elsewhere, has also been reported in single ultrafine particles in Houston (72) demonstrating unique Si sources in the industrialized ship channel region. Vegetative burning identified using K as a tracer and marine aerosols identified with Na and Mg are also important PM2.5 and PM10 sources in Houston (12, 20, 26, 47, 50). Interestingly, biomass burning contributed largely to fine particles whereas sea salt contributed to both PM2.5 and PM10. Source abundance profiles were developed for tailpipe and non-tailpipe PM2.5 and PM10 emissions from light-duty vehicles by making measurements inside the Washburn Tunnel (32). Elements such as Ca, Si, Fe, Al, Mg, K, and Ti displayed highest contributions especially in the coarse size mode. However, it would be nearly impossible to accurately isolate mobile source emissions using these elements since they are co-emitted in copious amounts via crustal resuspension. PGEs such as Rh, Pd, and Pt and other elements such as Cu, Zn, Ga, As, Zr, Mo, Cd, Sn, Sb, Ba, W, and Pb were prominently featured in the source profile. Enrichment factors for these elements in tunnel aerosols ranged between 3.5 for W and 36,575 for Rh in PM2.5 (see Table 2) (15, 32). Vehicular contributions to PM2.5 have been calculated to lie between 2.8 – 35.7% and between 0.8 – 30.0% 29 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

for PM10 mass (20, 26, 47) in Houston. Heavy duty vehicles were calculated to contribute to only ~5% of fine PM mass compared with 17.5% from light-duty vehicles (47). Marker elements signifying oil combustion and shipping activities include V, Ni, Sc, Mo (10, 12, 47). These metals have been measured in Houston aerosols apportioning oil combustion between 0.4 – 11% to PM2.5 and 0.2 – 35.0% to PM10 (10, 12, 20, 26, 47). Products of coal combustion and high temperature operations such as incineration, oil/coal fired boilers, smelters, and metal works such as Se, As, Cd, Sn, Sb, Pb, and Zn have also been identified (10, 12), apportioning these sources to 1.6 – 13.5% of PM2.5 and 0.6 – 9.6% of PM10 (10, 12, 20, 26, 50). As explained earlier, light lanthanoids including La, Ce, Pr, Nd, and Gd with minor contributions of Yb, Sm, Al, and Si has also been used to identify petroleum refining emissions (10, 12, 20, 26). Chemical mass balancing of samples collected in conjunction with cat-cracking emission events showed 12% contribution to PM2.520 whereas background refining contribution was only 0.6 – 2.1% (10, 12, 20, 26).

Concluding Remarks We summarized methodologies for digesting particulate matter and subsequent DRC-q-ICP-MS measurement of a suite of elements at trace – major levels. Long-term monitoring revealed the elemental signatures of several local and global sources contributing to ambient particulate matter levels in Houston. Specifically, measurements of lanthanoids and PGEs at trace levels are important to track primary emissions from petroleum refineries and light-duty gasoline-driven motor vehicles, respectively. Ambient aerosols in Houston are typically enriched in lanthanum and display significant differences in the lanthanoid abundance sequence compared with crustal material. This allows us to differentiate natural versus anthropogenic aerosol origins especially during non-routine operations of refinery catalytic cracking units. However, the lanthanoid composition of airborne particles was sometimes modified to closely resemble the crustal signature and substantially diluting the La-anomaly. These elemental signatures coupled with receptor modeling revealed periodic influences of trans-Atlantic dust transport from North Africa on aerosol mass concentrations in Texas. Concentrations of the second group of elements serving as source signatures, namely PGEs, were measured in road dust and airborne particles in a tunnel environment. A high level of care is necessary during sample preparation and q-ICP-MS to accurately and precisely measure Rh, Pd, and Pt, since these siderophilic (i.e. iron-loving) elements are present only at pg m–3 levels in the atmosphere. Being siderophiles, PGEs have dissolved in iron and migrated to the earth’s core thereby depleting themselves from the UCC. Highest enrichment of PGEs was measured inside the tunnel demonstrating them to be unique tracers for gasoline-driven automobile emissions. PGE composition of several United States matrices deviated substantially from corresponding European and Asian samples, emphasizing their geographical variability. Therefore, we recommend 30 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

PGE measurement in local samples for robust quantification of automobile contributions to measured aerosol mass. In closing, aerosols in industrialized urban environments originate from numerous local, regional, and global sources. Detailed information on individual sources and their contributions is necessary to better protect public health from particulate air pollution. One approach towards this goal that we have pursued over the years is to comprehensively measure the elemental composition of aerosols and use it as inputs to receptor models. This procedure identifies potential sources and quantifies their separate contributions to observed ambient PM2.5 and PM10 mass concentrations. It is emphasized that this method addresses only “primary” particles (i.e. those that are directly emitted) and explicitly ignores “secondary formation” (i.e. particles that are formed in the atmosphere from gaseous precursors). Nevertheless, such datasets can provide the scientific basis for the improvement of policies designed to protect public health associated with particulate air pollution. They also aid in improving public policy by focusing on relevant sources and in developing regulations based on knowledge of important sources that affect local ambient air quality.

Acknowledgments Portions of this work were funded by the Texas Commission on Environmental Quality and the Texas Air Research Center. We appreciate discussions with Profs. Matthew Fraser of Arizona State University and Joe Prospero of University of Miami during the course of this research.

References 1.

2.

3.

4.

5.

Pope, C. A.; Burnett, R. T.; Thun, M. J.; Calle, E. E.; Krewski, D.; Ito, K.; Thurston, G. D. Lung cancer, cardiopulmonary mortality, and long-term exposure to fine particulate air pollution. JAMA, J. Am. Med. Assoc. 2002, 287, 1132–1141. Chow, J. C.; Watson, J. G.; Kuhns, H.; Etyemezian, V.; Lowenthal, D. H.; Crow, D.; Kohl, S. D.; Engelbrecht, J. P.; Green, M. C. Source profiles for industrial, mobile, and area sources in the Big Bend Regional Aerosol Visibility and Observational study. Chemosphere 2004, 54, 185–208. Olmez, I.; Sheffield, A. E.; Gordon, G. E.; Houck, J. E.; Pritchett, L. C.; Cooper, J. A.; Dzubay, T. G.; Bennett, R. L. Compositions of particles from selected sources in Philadelphia for receptor modeling applications. J. Air Waste Manage. Assoc. 1988, 38, 1392–1402. Watson, J. G.; Chow, J. C.; Houck, J. E. PM2.5 chemical source profiles for vehicle exhaust, vegetative burning, geological material, and coal burning in Northwestern Colorado during 1995. Chemosphere 2001, 43, 1141–1151. Begum, B. A.; Hopke, P. K.; Zhao, W. Source Identification of Fine Particles in Washington, DC, By Expanded Factor Analysis Modeling. Environ. Sci. Technol. 2005, 39, 1129–1137. 31 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

6.

7. 8.

9.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

10.

11.

12.

13.

14.

15.

16.

17.

18. 19.

Lee, S.; Liu, W.; Wang, Y.; Russell, A. G.; Edgerton, E. S. Source apportionment of PM2.5: Comparing PMF and CMB results for four ambient monitoring sitesin the southeastern United States. Atmos. Environ. 2008, 42, 4126–4137. Johnson, J.; Schewel, L.; Graedel, T. E. The Contemporary Anthropogenic Chromium Cycle. Environ. Sci. Technol. 2006, 40, 7060–7069. Hjortenkrans, D. S. T.; Bergback, B. G.; Haggerud, A. V. Metal Emissions from Brake Linings and Tires: Case Studies of Stockholm, Sweden 1995/ 1998 and 2005. Environ. Sci. Technol. 2007, 41, 5224–5230. Shah, P.; Strezov, V.; Prince, K.; Nelson, P. F. Speciation of As, Cr, Se and Hg under coal fired power station conditions. Fuel 2008, 87, 1859–1869. Bozlaker, A.; Buzcu-Guven, B.; Fraser, M. P.; Chellam, S. Insights into PM10 sources in Houston, Texas: Role of petroleum refineries in enriching lanthanoid metals during episodic emission events. Atmos. Environ. 2013, 69, 109–117. Kitto, M. E.; Anderson, D. L.; Gordon, G. E.; Olmez, I. Rare Earth Distribution in Catalysts and Airborne Particles. Environ. Sci. Technol. 1992, 26, 1368–1375. Kulkarni, P.; Chellam, S.; Fraser, M. P. Lanthanum and lanthanides in atmospheric fine particles and their apportionment to refinery and petrochemical operations in Houston, TX. Atmos. Environ. 2006, 40, 508–520. Moreno, T.; Querol, X.; Alastuey, A.; Pey, J.; Minguillon, M. C.; Perez, N.; Bernabe, R. M.; Blanco, S.; Cardenas, B.; Gibbons, W. Lanthanoid Geochemistry of Urban Atmospheric Particulate Matter. Environ. Sci. Technol. 2008, 42, 6502–6507. Rauch, S.; Peucker-Ehrenbrink, B.; Molina, L. T.; Molina, M. J.; Ramos, R.; Hemond, H. F. Platinum Group Elements in Airborne Particles in Mexico City. Environ. Sci. Technol. 2006, 40, 7554–7560. Spada, N. J.; Bozlaker, A.; Chellam, S. Multi-elemental characterization of tunnel and road dusts in Houston, texas using dynamic reaction cell-quadrupole-inductively coupled plasma-mass spectrometry: Evidence for the release of platinum group and anthropogenic metals from motor vehicles. Anal. Chim. Acta 2012, 735, 1–8. Zereini, F.; Wiseman, C.; Puettmann, W. Changes in Palladium, Platinum, And Rhodium Concentrations, And Their Spatial Distribution in Soils along a Major Highway in Germany from 1994 to 2004. Environ. Sci. Technol. 2007, 41, 451–456. Migaszewski, Z. M.; Galuszka, A. The characteristics, occurrence, and geochemical behavior of rare earth elements in the environment: A review. Crit. Rev. Environ. Sci. Technol. 2015, 45, 429–471. Olmez, I.; Gordon, G. E. Rare earths: Atmospheric signatures for oil-fired power plants and refineries. Science 1985, 229, 966–968. Danadurai, K. S. K.; Chellam, S.; Lee, C. T.; Fraser, M. P. Trace elemental analysis of airborne particulate matter using dynamic reaction cell inductively coupled plasma - mass spectrometry: Application to monitoring episodic industrial emission events. Anal. Chim. Acta 2011, 686, 40–49. 32 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

20. Kulkarni, P.; Chellam, S.; Fraser, M. P. Tracking Petroleum Refinery Emission Events Using Lanthanum and Lanthanides As Elemental Markers For PM2.5. Environ. Sci. Technol. 2007, 41, 6748–6754. 21. Kitto, M. E. Trace-element patterns in fuel oils and gasolines for use in source apportionment. J. Air Waste Manage. Assoc. 1993, 43, 1381–1388. 22. Goudie, A. S.; Middleton, N. J. Saharan dust storms: nature and consequences. Earth-Sci. Rev. 2001, 56, 179–204. 23. Goudie, A. S.; Middleton, N. J. Desert Dust in the Global System; Springer: Heidelberg, Germany, 2006. 24. Prospero, J. M.; Ginoux, P.; Torres, O.; Nicholson, S. E.; Gill, T. E. Environmental characterization of global sources of atmospheric soil dust identified with the Nimbus 7 Total Ozone Mapping Spectrometer (TOMS) absorbing aerosol product. Rev. Geophys. 2002, 40. 25. Prospero, J. M. African dust in America. Geotimes 2001, 46, 24–27. 26. Bozlaker, A.; Prospero, J. M.; Fraser, M. P.; Chellam, S. Quantifying the Contribution of Long-Range Saharan Dust Transport on Particulate Matter Concentrations in Houston, Texas, Using Detailed Elemental Analysis. Environ. Sci. Technol. 2013, 47, 10179–10187. 27. Negral, L.; Moreno-Grau, S.; Moreno, J.; Querol, X.; Viana, M. M.; Alastuey, A. Natural and anthropogenic contributions to PM10 and PM2.5 in an urban area in the western Mediterranean coast. Water, Air, Soil Pollut. 2008, 192, 227–238. 28. Rauch, S.; Peucker-Ehrenbrink, B., Sources of Platinum Group Elements in the Environment. In Platinum Metals in the Environment; Zereini, F., Wiseman, C., Eds.; Springer: Heidelberg, Germany, 2015; pp 199−242. 29. Konig, H. P.; Hertel, R. F.; Koch, W.; Rosner, G. Determination of platinum emissions from a three-way catalyst-equipped gasoline engine. Atmos. Environ. 1992, 26A, 741–745. 30. Zereini, F.; Wiseman, C. L. S. Platinum Metals in the Environment; Springer: Berlin Heidelberg, 2015. 31. Rauch, S.; Hemond, H. F.; Barbante, C.; Owari, M.; Morrison, G. M.; Peucker-Ehrenbrink, B.; Wass, U. Importance of Automobile Exhaust Catalyst Emissions for the Deposition of Platinum, Palladium, And Rhodium in the Northern Hemisphere. Environ. Sci. Technol. 2005, 39, 8156–8162. 32. Bozlaker, A.; Spada, N. J.; Fraser, M.; Chellam, S. Elemental characterization of PM2.5 and PM10 emitted from light duty vehicles in the Washburn Tunnel of Houston, TX: Release of rhodium, palladium, and platinum. Environ. Sci. Technol. 2014, 48, 54–62. 33. Chellam, S.; Bozlaker, A., Characterization of PGEs and other elements in road dusts and airborne particles in Houston, Texas. In Platinum Metals in the Environment; Zereini, F., Wiseman, C. L. S., Eds.; Springer: Berlin Heidelberg, 2015; pp 199−242. 34. Ely, J. C.; Neal, C. R.; Kulpa, C. F.; Schneegurt, M. A.; Seidler, J. A.; Jain, J. C. Implications of Platinum-Group Element Accumulation along U.S. Roads from Catalytic-Converter Attrition. Environ. Sci. Technol. 2001, 35, 3816–3822. 33 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

35. Moldovan, M. Origin and fate of platinum group elements in the environment. Anal. Bioanal. Chem. 2007, 388, 537–540. 36. Moldovan, M.; Palacios, M. A.; Gomez, M. M.; Morrison, G.; Rauch, S.; McLeod, C.; Ma, R.; Caroli, S.; Alimonti, A.; Petrucci, F.; Bocca, B.; Schramel, P.; Zischka, M.; Pettersson, C.; Wass, U.; Luna, M.; Saenz, J. C.; Santamaria, J. Environmental risk of particulate and soluble platinum group elements released from gasoline and diesel engine catalytic converters. Sci. Total Environ. 2002, 296, 199–208. 37. U.S. Department of Transportation, Federal Highway Administration. http://www.fhwa.dot.gov/policyinformation/travel_monitoring/tvt.cfm (accessed on September 24, 2015). 38. Kulkarni, P.; Chellam, S.; Flanagan, J. B.; Jayanty, R. K. M. Microwave digestion - ICP-MS for elemental analysis in ambient airborne fine particulate matter: Rare earth elements and validation using a filter borne fine particle certified reference material. Anal. Chim. Acta 2007, 599, 170–176. 39. Kulkarni, P.; Chellam, S.; Mittlefehldt, D. W. Microwave-assisted extraction of rare earth elements from petroleum refining catalysts and ambient fine aerosols prior to inductively coupled plasma-mass spectrometry. Anal. Chim. Acta 2007, 581, 247–259. 40. D’Ilio, S.; Violante, N.; Majorani, C.; Petrucci, F. Dynamic reaction cell ICPMS for determination of total As, Cr, Se and V in complex matrices: Still a challenge? A review. Anal. Chim. Acta 2011, 698, 6–13. 41. Djingova, R.; Heidenreich, H.; Kovacheva, P.; Markert, B. On the determination of platinum group elements in environmental materials by inductively coupled plasma mass spectrometry and microwave digestion. Anal. Chim. Acta 2003, 489, 245–251. 42. Bencs, L.; Ravindra, K.; Van Grieken, R. Methods for the determination of platinum group elements originating from the abrasion of automotive catalytic converters. Spectrochim. Acta, Part B 2003, 58, 1723–1755. 43. Borisov, O. V.; Coleman, D. M.; Oudsema, K. A.; Carter, R. O. Determination of platinum, palladium, rhodium and titanium in automotive catalytic converters using inductively coupled plasma mass spectrometry with liquid nebulization. J. Anal. At. Spectrom. 1997, 12, 239–246. 44. Lesniewska, B. A.; Godewska-Zylkiewicz, B.; Bocca, B.; Caimi, S.; Caroli, S.; Hulanicki, A. Platinum, palladium and rhodium content in road dust, tunnel dust and common grass in Bialystok area (Poland): a pilot study. Sci. Total Environ. 2004, 321, 93–104. 45. Lesniewska, B. A.; Godlewska-Zylkiewicz, B.; Ruszczynska, A.; Bulska, E.; Hulanicki, A. Elimination of interferences in determination of platinum and palladium in environmental samples by inductively coupled plasma mass spectrometry. Anal. Chim. Acta 2006, 564, 236–242. 46. Fraser, M. P.; Yue, Z. W.; Trapp, J. M.; Kohl, S. D.; Chow, J. C. Molecular composition of organic fine particulate matter in Houston, TX. Atmos. Environ. 2002, 36, 5751–5758. 47. Sullivan, D. W.; Price, J. H.; Lambeth, B.; Sheedy, K. A.; Savanich, K.; Tropp, R. J. Field study and source attribution for PM2.5 and PM10 with resulting reduction in concentrations in the neighborhood north of the 34 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

48.

49.

50.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

51.

52.

53. 54.

55.

56.

57.

58.

59.

60.

Houston Ship Channel based on voluntary efforts. J. Air Waste Manage. Assoc. 2013, 63, 1070–1082. Chellam, S.; Kulkarni, P.; Fraser, M. P. Emissions of organic compounds and trace metals in fine particulate matter from motor vehicles: A tunnel study in Houston, Texas. J. Air Waste Manage. Assoc. 2005, 55, 60–72. Watson, J. G.; Robinson, N. F.; Chow, J. C.; Henry, R. C.; Kim, B. M.; Pace, T. G.; Meyer, E. L.; Nguyen, Q. The USEPA/DRI chemical mass balance receptor model, CMB 7.0. Environ. Softw. 1990, 5, 38–49. Buzcu, B.; Fraser, M. P.; Kulkarni, P.; Chellam, S. Source identification and apportionment of fine particulate matter in Houston, TX, using positive matrix factorization. Environ. Eng. Sci. 2003, 20, 533–545. Dillner, A. M.; Schauer, J. J.; Christensen, W. F.; Cass, G. R. A quantitative method for clustering size distributions of elements. Atmos. Environ. 2005, 39, 1525–1537. Park, J. W.; Hu, Z. C.; Gao, S.; Campbell, I. H.; Gong, H. J. Platinum group element abundances in the upper continental crust revisited - New constraints from analyses of Chinese loess. Geochim. Cosmochim. Acta 2012, 93, 63–76. Wedepohl, K. H. The composition of the continental crust. Geochim. Cosmochim. Acta 1995, 59, 1217–1232. Du, L.; Turner, J. Using PM2.5 lanthanoid elements and nonparametric wind regression totrac k petroleum refinery FCC emissions. Sci. Total Environ. 2015, 529, 65–71. Moreno, T.; Querol, X.; Alastuey, A.; Gibbons, W. Identification of FCC refinery atmospheric pollution events using lanthanoid- and vanadium-bearing aerosols. Atmos. Environ. 2008, 42, 7851–7861. Moreno, T.; Querol, X.; Alastuey, A.; de la Rosa, J.; de la Campa, A. M. S.; Minguillon, M.; Pandolfi, M.; Gonzalez-Castanedo, Y.; Monfort, E.; Gibbons, W. Variations in vanadium, nickel and lanthanoid element concentrations in urban air. Sci. Total Environ. 2010, 408, 4569–4579. de la Campa, A. M. S.; Moreno, T.; de la Rosa, J.; Alastuey, A.; Querol, X. Size distribution and chemical composition of metalliferous stack emissions in the San Roque petroleum refinery complex, southern Spain. J. Hazard. Mater. 2011, 190, 713–722. Moreno, T.; Perez, N.; Querol, X.; Amato, F.; Alastuey, A.; Bhatia, R.; Spiro, B.; Hanvey, M.; Gibbons, W. Physicochemical variations in atmospheric aerosols recorded at sea onboard the Atlantic-Mediterranean 2008 Scholar Ship cruise (Part II): Natural versus anthropogenic influences revealed by PM10 trace element geochemistry. Atmos. Environ. 2010, 44, 2563–2576. Moreno, T.; Alastuey, A.; Querol, X.; Font, O.; Gibbons, W. The identification of metallic elements in airborne particulate matter derived from fossil fuels at Puertollano, Spain. Int. J. Coal Geol. 2007, 71, 122–128. Jarvis, K. E.; Parry, S. J.; Piper, J. M. Temporal and Spatial Studies of Autocatalyst-Derived Platinum, Rhodium, And Palladium and Selected Vehicle Derived Trace Elements in the Environment. Environ. Sci. Technol. 2001, 35, 1031–1036. 35 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch001

61. Rauch, S.; Hemond, H. F.; Peucker-Ehrenbrink, B.; Ek, K. H.; Morrison, G. M. Platinum Group Element Concentrations and Osmium Isotopic Composition in Urban Airborne Particles from Boston, Massachusetts. Environ. Sci. Technol. 2005, 39, 9464–9470. 62. Sutherland, R. A.; Pearson, D. G.; Ottley, C. J. Platinum-group elements (Ir, Pd, Pt and Rh) in road-deposited sediments in two urban watersheds, Hawaii. Appl. Geochem. 2007, 22, 1485–1501. 63. Sutherland, R. A.; Pearson, D. G.; Ottley, C. J. Grain size partitioning of platinum-group elements in road-deposited sediments: Implications for anthropogenic flux estimates from autocatalysts. Environ. Pollut. 2008, 151, 503–515. 64. Gomez, B.; Palacios, M. A.; Gomez, M.; Sanchez, J. L.; Morrison, G.; Rauch, S.; McLeod, C.; Ma, R.; Caroli, S.; Alimonti, A.; Petrucci, F.; Bocca, B.; Schramel, P.; Zischka, M.; Petterson, C.; Wass, U. Levels and risk assessment for humans and ecosystems of platinum-group elements in the airborne particles and road dust of some European cities. Sci. Total Environ. 2002, 299, 1–19. 65. Gomez, M. B.; Gomez, M. M.; Palacios, M. A. ICP-MS determination of Pt, Pd and Rh in airborne and road dust after tellurium coprecipitation. J. Anal. At. Spectrom. 2003, 18, 80–83. 66. Kanitsar, K.; Koellensperger, G.; Hann, S.; Limbeck, A.; Puxbaum, H.; Stingeder, G. Determination of Pt, Pd and Rh by inductively coupled plasma sector field mass spectrometry (ICP-SFMS) in size-classified urban aerosol samples. J. Anal. At. Spectrom. 2003, 18, 239–246. 67. Petrucci, F.; Bocca, B.; Alimonti, A.; Caroli, S. Determination of Pd, Pt and Rh in airborne particulate and road dust by high-resolution ICP-MS: a preliminary investigation of the emission from automotive catalysts in the urban area of Rome. J. Anal. At. Spectrom. 2000, 15, 525–528. 68. Rauch, S.; Lu, M.; Morrison, G. M. Heterogeneity of Platinum Group Metals in Airborne Particles. Environ. Sci. Technol. 2001, 35, 595–599. 69. Zereini, F.; Alt, F.; Messerschmidt, J.; Von Bohlen, A.; Liebl, K.; Puttmann, W. Concentration and Distribution of Platinum Group Elements (Pt, Pd, Rh) in Airborne Particulate Matter in Frankfurt am Main, Germany. Environ. Sci. Technol. 2004, 38, 1686–1692. 70. Zereini, F.; Wiseman, C. L. S.; Puttmann, W. In Vitro Investigations of Platinum, Palladium, and Rhodium Mobility in Urban Airborne Particulate Matter (PM10, PM2.5, and PM1) Using Simulated Lung Fluids. Environ. Sci. Technol. 2012, 46, 10326–10333. 71. Zereini, F.; Alsenz, H.; Wiseman, C. L. S.; Puttmann, W.; Reimer, E.; Schleyer, R.; Bieber, E.; Wallasch, M. Platinum group elements (Pt, Pd, Rh) in airborne particulate matter in rural vs. urban areas of Germany: Concentrations and spatial patterns of distribution. Sci. Total Environ. 2012, 416, 261–268. 72. Phares, D. J.; Rhoads, K. P.; Johnston, M. V.; Wexler, A. S. Size-resolved ultrafine particle composition analysis - 2. Houston. J. Geophys. Res.: Atmos. 2003, 108. 36 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Chapter 2

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

A Review of Ozone Studies in the Houston−Galveston−Brazoria Nonattainment Area Md. Tarkik Shahriar,1,2 Akhil Kadiyala,1 Raghava Kommalapati,*,1,2 and Ziaul Huque1,3 1Center for Energy & Environmental Sustainability, Prairie View A&M University, Prairie View, Texas 77446, U.S.A. 2Department of Civil & Environmental Engineering, Prairie View A&M University, Prairie View, Texas 77446, U.S.A. 3Department of Mechanical Engineering, Prairie View A&M University, Prairie View, Texas 77446, U.S.A. *E-mail: [email protected].

The United States Environmental Protection Agency identified the Houston-Galveston-Brazoria (HGB) area as a major nonattainment area for ozone (O3) exceedances. Several studies have been reported in the literature that examined the various factors facilitating the O3 exceedances in the HGB area. However, there was no consolidation of research studies representing the HGB area O3 exceedances. This book chapter provides a summary of the findings from research studies associated with O3 exceedances in the HGB area. Based on the literature review, this book chapter noted how the emission of highly reactive volatile organics and nitrogen oxide from industrial facilities in combination with favorable meteorological conditions (e.g., sea breeze circulation, higher solar radiation, weak winds) significantly contributed to O3 exceedances.

© 2015 American Chemical Society In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

Introduction Ozone (O3) is a secondary air pollutant formed through a series of atmospheric chemical reactions in the troposphere. The tropospheric O3 (also referred to as ground-level O3) is formed when two of the precursors, namely, volatile organic compounds (VOCs) and nitrogen oxide (NOx) go through a series of photochemical reactions (1–5). The O3 formation mechanism in troposphere includes the following set of chemical reactions (6). First, reactive VOCs (RH) interact with hydroxyl (OH) radicals to form organic radicals ( ) (reaction 1). Next, the organic radicals combine with molecular oxygen (O2) to form peroxy radicals ( ) (reaction 2). then reacts with nitric oxide (NO) to form nitrogen dioxide (NO2) (reaction 3). In the presence of solar radiation (hv), NO2 is further broken down to form NO and an oxygen atom (O) (reaction 4). Finally, O2 combines with O to form O3 (reaction 5). The O3 formation is a chain reaction process driven by the resulting OH radical formation. The newly formed O3 obtains energy from ultraviolet hv resulting in the development of O (reaction 6). O then reacts with water vapor (H2O) to form two OH radicals (reaction 7). Moreover, further production of OH radicals is possible in the presence of NO, which can be initiated by the RO radical formation (reaction 3). Availability of sufficient VOCs and NOx results in the formation and accumulation of tropospheric O3 as shown by reactions 1 through 7.

Ozone impacts both human health and vegetation. Several studies confirmed a positive relationship between elevated O3 concentrations and premature mortality (7–13). The intake of ground-level O3 can lead to health problems resulting in cough, chest pain, throat congestion and irritation, reduced lung function, and inflammation of linings of the lungs in addition to worsening the conditions of asthma, bronchitis, and emphysema. Agricultural yields of wheat (14, 15), sugar beet (16), potatoes (17), and oilseed rape (18, 19) were observed to be reduced by the phytotoxic properties of O3. Experimental studies have shown 38 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

that the growth and yield of crops (20–23) can be highly affected by existing O3 levels. Several studies (24–32) noted short-term high O3 concentration exposures to be more potent than long-term low O3 concentration exposures with respect to vegetation. The United States Environmental Protection Agency (USEPA) in accordance with the 1990 Clean Air Act set National Ambient Air Quality Standards (NAAQS) for O3 to ensure the safety of public health. The 1990 1-h O3 NAAQS was set to 125 parts per billion (ppb) in accordance with an earlier legislated February 8, 1979 Rule 44 FR 8202. The inability of the 1990 1-h O3 NAAQS to address public health concerns led to the emergence of a more stringent 8-h O3 NAAQS in 1997. The 1997 8-h O3 NAAQS was set to 84 parts per billion (ppb). To determine the O3 attainment status of a particular region, a design value calculated by averaging the fourth-highest 8-h average O3 concentrations across three consecutive years was proposed under the 8-h O3 NAAQS by USEPA. A revision of the 8-h O3 NAAQS was made on March 23, 2008 by reducing the limit from 84 ppb to 75 ppb. The NAAQS are re-evaluated every five years by the USEPA on the basis of best available science to ensure adequate protection of human health. The next revision for 8-h O3 NAAQS, which is currently under review, is proposed to set the limit between 65 ppb and 70 ppb and anticipated to be published as a final rule in October 2015 (33). The Houston-Galveston-Brazoria (HGB) area has been one of the O3 nonattainment areas in the U.S.A. This book chapter is aimed at reviewing existing O3 air quality studies to assess the scope for improvements in developing air quality modeling tools and providing the public with a consolidated summary of O3 studies specifically in the HGB area.

Houston−Galveston−Brazoria Ozone Nonattainment Study Area Figure 1 illustrates the 8-h O3 nonattainment areas based on 2008 NAAQS. A nonattainment area is defined as an area having air pollutant levels persistently exceeding the designated NAAQS levels. The HGB area, geographically located in southeast Texas, consists of eight counties - Brazoria, Chambers, Fort Bend, Galveston, Harris, Liberty, Montgomery, and Waller (refer Figure 2). The HGB area was designated as a severe O3 nonattainment area under the 1997 8-h O3 NAAQS and categorized as a marginal O3 nonattainment area under the 2008 8-h O3 NAAQS. Some studies examined the O3 formation mechanism in the HGB area using photochemical models that simulate air quality based on a set of mathematical equations characterizing the chemical and physical processes. The Comprehensive Air Quality Model with extensions (CAMx) and the Community Multi-scale Air Quality (CMAQ) model are two such photochemical modeling tools adopted by the researchers and regulatory agencies in examining the HGB area O3 formation. Figure 3 illustrates the boundary domain for the HGB area CAMx photochemical modeling grid as specified by Texas Commission on Environmental Quality (TCEQ). The HGB-Beaumont-Port Arthur (HGBPA) modeling subdomain (represented by light blue-colored boundary in Figure 3) has a range (km) with 39 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

(431,505) easting and (-1153,-1079) northing, and is represented by 74 × 74 cells with each cell sizing 1 km × 1 km, within the East Texas Subdomain (represented by green-colored boundary in Figure 3) of the Regional domain (represented by dark blue-colored boundary in Figure 3).

Figure 1. O3 nonattainment areas in the U.S.A. based on 2008 NAAQS. HGB area is represented in blue in southeast Texas. Reproduced with permission from Ref. (34). Copyright (2015) USEPA. (see color insert)

Figure 2. Eight counties in the HGB area. Reproduced with permission from Ref. (35). Copyright (2014) Texas Commission on Environmental Quality. (see color insert) 40 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

Figure 3. CAMx modeling domain grid with HGBPA modeling subdomain (represented by light blue-colored boundary) within the East Texas Subdomain (represented by green-colored boundary) of the Regional domain (represented by dark blue-colored boundary). Reproduced with permission from Ref. (36). Copyright (2014) Texas Commission on Environmental Quality. (see color insert)

Review of Ozone Air Quality Studies in Houston−Galveston−Brazoria Area Daum et al. (37) evaluated and compared the O3 formation rates in two distinct areas within the HGB area: downtown Houston and the Houston Ship Channel. The O3 formation rates in downtown Houston (3-18 ppb/h) were much lower than the O3 formation rates in the Houston Ship Channel (3-80 ppb/h). The higher O3 formation rates in the Houston Ship Channel area are a result of much higher hydrocarbon reactivity (mainly comprising of low molecular weight alkenes). The flight-monitored O3 observations in downtown Houston on August 29, 2000 changed from 30 ppb in the morning to 65 ppb in the early afternoon, and 80 ppb in the late afternoon. The O3 formation in downtown Houston during 41 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

the mornings is driven by point sources such as the Parrish Power Plant located to the southwest of the city. The influence of point sources decreased as the day progressed in downtown Houston. The flight-monitored O3 observations in the Houston Ship Channel on August 29, 2000 were noted to be consistently higher than the O3 observations in downtown Houston by approximately 80 ppb throughout the day and reached a maximum of 180 ppb in the late afternoon. The higher O3 concentrations in the Houston Ship Channel were noted to be a result of a combination of plumes from separate sources. O3 formation in downtown Houston was noted to be much more hydrocarbon limited than in the Houston Ship Channel. Zhou et al. (38) compared the flight-monitored O3 observations in Houston area between 2000 and 2006 summer (August–October) episodes. The averaged O3 concentrations in 2000 and 2006 were noted to be 76 ppb and 58 ppb, respectively. The decrease in O3 concentrations between 2000 and 2006 may be attributed to the implementation of major emission reduction measures undertaken by petrochemical and other sources. The reduction in NOx and highly reactive VOCs (HRVOCs) largely contributed a 40-50% reduction in O3 formation rates from 2000 to 2006. The reduction in HRVOCs also decreased the radical production by 20-50%, further inhibiting the O3 formation rates in Houston. Murphy and Allen (39) extensively studied the hydrocarbon emissions from industrial release events and their associated impact on O3 formation in the Houston-Galveston (HG) area. The “events” or “upsets” were defined as non-routine emissions where the released quantities of those emissions are greater than normal and where such phenomena are observed for a shorter period of time (typically less than 24 h). The online event reporting system, maintained by TCEQ, was used to examine the magnitudes and variability of emission events between January 31, 2003 and January 30, 2004. Flow rates (lb/h) of NOX, VOCs, and HRVOCs were calculated to facilitate computing the magnitude and frequency of those emission events, which were then compared against annual average flow rates within the region and at specific facilities. They identified O3 formation in the HG area to be significantly influenced by the HRVOC emissions and NOX emissions having negligible influence. More than half of the mass of HRVOC event emissions were characterized as ethene, one-third as propene, and the remaining 10% as isomers of butene and 1,3-butadiene. Ethene also dominated the frequency events. There were 761 HRVOC events during the study period, among which 87 (11%) lasted 10 min or less, 192 (25%) lasted 1 h or less, and 567 (75%) lasted 24 h or less. Although the duration of the events was short, their magnitudes were significantly large. In the case of HRVOC events lasting 1 h or less, 33 of the events were reported to release more than 1000 lbs of HRVOCs. Industrial Organic Chemicals were identified to be the primary components accounting for two-thirds of the mass of HRVOC emission events and 90% of the mass in a 12-mo reporting period was attributed to about 20 facilities. Li et al. (40) comprehensively analyzed the impacts of biogenic emissions on photochemical O3 production in the Houston area by applying a three-dimensional regional chemical transport model (CTM) using the Texas Air Quality Study 2000 (TexAQS 2000) database. The simulated O3 concentrations were compared with 42 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

observations from the G-1 research aircraft and a surface-monitoring network. The CTM model performed well in reproducing the location, magnitude, and movements of O3 plumes along with the temporal and geographical variations of O3. The simulated concentrations of total reactive nitrogen (NOy) and the variations of peroxyacetyl nitrate (PAN) from the CTM were in agreement with the monitored observations. Despite the uncertainties that existed in the case of isoprene (an important biogenic hydrocarbon) emissions and meteorological inputs, the CTM simulations compared reasonably with available observations. O3 plume was observed to occur in the urban Houston area in the afternoon, with isoprene emissions playing a major role accounting for approximately 20–40 ppb of O3 concentrations in the Houston area. The study also analyzed the O3 formation by examining the VOC/NO2 reactivity ratios. The VOC/NO2 reactivity ratios were relatively higher in the Houston Ship Channel than in the urban Houston area. This pattern was observed to be opposite for the case of isoprene emissions. When the isoprene emissions were decreased or increased by 50%, the O3 concentrations changed proportionately by 5–25 ppb over the urban Houston area, while the change in the O3 concentration was less than 5–10 ppb over the Houston Ship Channel. Ryerson et al. (41) analyzed the O3 formation potential of petrochemical industrial emissions of reactive alkenes and NOx in the Houston area using two days of TexAQS 2000 database. A three-step approach was adopted. First, the plumes of some geographically isolated complexes (at Freeport, Chocolate Bayou, and Sweeny) were analyzed to evaluate their O3 production potential. Next, data from combined plumes downwind of multiple petrochemical complexes in Texas City area and heavily industrialized Houston Ship Channel were included to study their impact on O3 production potential. Finally, a comparison was made between O3 productions from petrochemical industrial plumes against the observed downwind concentrations of fossil-fueled electric utility power plants in urban/rural areas. Petrochemical emissions’ contribution to O3 formation rates and yields are substantially higher than the contributions of fossil-fueled electric utility power plant emissions. All possible VOCs that are emitted from petrochemical sources were identified to be utilized in O3 production in Houston area. Accurate estimation of reactive light alkenes was noted to be essential in the development of exact VOC emission inventories for Houston petrochemical industrial sources. By decreasing the emissions of ethene and propene, a reduction of over 75% of initial VOC reactivity was observed to have been possible, thereby indicating a reduction in alkene emissions is essential to control O3 exceedances in the Houston area. Washenfelder et al. (42) characterized NOx, sulfur dioxide (SO2), ethene, and propene from industrial emission sources in the Houston area by examining the trends of emissions from industrial sources between 2000 and 2006. HRVOCs were additionally identified using relative OH radical reactivity. The use of abatement controls at industrial facilities provided a reduction of 29% ± 20% (mean ± standard deviation) in NOx emissions between 2000 and 2006. A reduction of 30% ± 30% in emissions of alkenes (ethene, propene) was also observed during the same period. These statistics were based on the examination of temporal trends in ethene/NOx and propene/NOx ratios from 43 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

isolated petrochemical sources. Within the Houston urban area between 2000 and 2006, the median ambient concentrations of ethene and propene have been reduced by 52% and 48%, respectively. The study also identified the continuation of inaccuracies in emission inventory for the 2000, 2002, and 2006 measurements in the HGB area. The emission inventory values exceeded the measured ratios of ethene/NOx and propene/NOx by factors of 1.4–20 and 1–24, respectively. Inaccurate accounting of ethene, propene, and other reactive VOC emissions may have resulted in the difficulty for CTMs to accurately predict O3 levels in the greater Houston area. The study noted the Houston Ship Channel to be the largest source among the many other petrochemical industrial areas (Texas City, Chocolate Bayou, Sweeny, and Freeport) examined. Webster et al. (43) examined the influence of variability in continuous hydrocarbon emissions on O3 formation in the HG area. The methodology involved (a) obtaining a set of observations from various flares and cooling towers and (b) developing models to simulate emission variability using the process of stochastic emissions inventory generation. The impact of industrial point source variability on O3 formation was assessed using CAMx for two episode days in August 2000. A total of 50 sets of stochastic emission inventories were randomly generated with the models for each day resulting in a total of 100 simulations. The maximum increase and decrease in peak O3 concentration was noted to be approximately 10.7 ppb and 4.5 ppb, respectively. These results indicate that the variability in continuous industrial emissions has a significant impact on O3 formation in the HG area. The variability in continuous industrial emissions addresses the variation of O3 concentrations in the HG area by approximately 10–52 ppb. Neuman et al. (44) examined the dependence of photochemical O3 production on oxidation of one of its precursors, NOx, in Houston by evaluating the plumes from the same source region under a variety of meteorological conditions. The study proposed the use of O3 production efficiency (OPE) to calculate the number of O3 molecules formed for each NOx molecule oxidized in analyzing the relationship between photochemical O3 production and NOx oxidation. OPE was calculated by accounting for the influence of changing backgrounds on the measured mixing ratios. Fast response (typically 1 Hz) measurements of PANs, NO2, NO, O3, NOy, carbon monoxide (CO) and nitric acid (HNO3) were analyzed. In plumes downwind from Houston industrial and urban areas, the observed enhancement ratios of ΔO3 and Δ(NOy-NOx) were used to determine OPE. Transport and mixing of pollutants was found to be quite complicated on six daytime flights in Houston. At different times and locations, the plumes were observed to be mixed with the background air pollutants, eventually resulting in varying background concentrations. With an increase in downwind distance, the ratios of ΔCO/ΔNOy and ΔO3/Δ(NOy-NOx) increased. Ratio of ΔO3/Δ(NOy-NOx) was also found to be highly variable and elevated with the change of background concentrations and increment of ΔCO/ΔNOy downwind. Rapid formation of O3 and PANs as well as higher OPE were observed in plumes dominated by Houston Ship Channel emissions. The ratio of O3 to NOx oxidation products ranged between 11 and 12 in the plumes heavily influenced by the Houston Ship Channel as also observed in other studies (41, 45). 44 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

Sarker et al. (46) analyzed the influence of point source VOCs on the HGB air quality by examining the O3 sensitivity with the ratio of hydrogen peroxides (H2O2) to nitric acid (HNO3). The H2O2/HNO3 ratios showed that O3 formation in the HGB area was NOx-limited for the simulated episode, thereby indicating the control of NOx emissions is imperative to control O3 exceedances in the HGB area. VOCs such as acetaldehyde (CH3CHO), formaldehyde (CH2O), ethane (C2H6) and PAN showed good correlation (R2) of 0.81, 0.53, 0.5, and 0.67, respectively with high O3 concentrations. Sulfur dioxide (SO2) had good correlation with VOCs, such as CH3CHO and C2H6 indicating that the reactivity of specific VOC species is significant for O3 exceedances in the HGB area. Xiao et al. (47) noted the HGB O3 formation to be highly nonlinear and dependent on the precursors (NOx, VOC) originating from different emission source categories. Petrochemical facilities in the Houston Ship Channel area were identified as the major sources for both NOx and HRVOC emissions that facilitate rapid and effective O3 formation. The characterized VOC emissions density for different species (CH3CHO, ethene, CH2O, isoprene, olefin carbon bond, paraffin carbon bond, toluene, and xylene) from the Houston Ship Channel and the rest of HGB were noted to be (11.9, 20, 12.5, 0.2, 18.1, 220.9, 18.4, 11.6) and (0.8, 0.8, 0.4, 0.05, 0.9, 13.1, 2.5, 1.9) tons/d/1000 km2, respectively. The CMAQ model sensitivity analysis revealed the Houston Ship Channel petrochemical facilities to be the larger contributors to peak O3 concentrations in the Houston region. NOx emission releases from the Houston Ship Channel (100 tons/d) and other local region sources were also noted to significantly influence the daily peak O3 concentrations. Byun et al. (48) examined the performance of CAMx and CMAQ photochemical models in the case of a high O3 event in the HGB area using same emissions and meteorological data as inputs. The two model simulation results were compared against the aircraft measurements from the TexAQS 2000 study. CMAQ was more assisted by the imputed HRVOC emissions in simulating observed peak O3 concentrations in comparison to CAMx. In some of the highly HRVOC-rich areas, such as downwind of the Houston Ship Channel and other surrounding areas, the O3 peaks predicted by CAMx were found to be higher than those predicted by CMAQ with imputed HRVOC emissions. The study noted the performance of CMAQ system to be essentially poor based on the results of base and imputed case simulations. Sarker et al. (49) analyzed the performance of different chemical mechanisms (CB5, CB6) using CAMx in predicting O3 concentrations in the HGB area. The CAMx CB5 and CB6 results were compared with the observations from TCEQ monitoring stations in the HGB. CB6 chemical mechanism provided better prediction of O3 concentrations than CB5. CB5 O3 predictions were approximately 15% lower than CB6 predictions. Though CB6 improved air quality prediction capacity for CAMx, further modifications are required to accurately predict O3 concentrations. Several studies strived to analyze in detail the influence of meteorology and geographical locations on the HGB area O3 exceedances (42, 50–54). The high levels of O3 in the HGB area are accredited to anthropogenic NOx and VOC sources, which, during the day time of intense solar radiation and stagnant 45 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

meteorological conditions, combine to form O3 (42). Banta et al. (50) noted that the meteorology of the HGB is favorable for O3 production on the basis of sea breeze circulation that confines the pollutants to the urban area, resulting in accumulation of high levels of O3. Nielson-Gammon et al. (51) noted the months of August and September to be the most susceptible times of the year for high O3 exceedances due to favorable meteorological conditions. Abundant sunshine, high temperatures, weak local winds, and higher frequency of winds from the north during September play a major role in increasing the background O3 in Houston in this time of the year (52). Being located close to the Gulf of Mexico, the HGB is one of the largest Metropolitan areas in the United States that often experiences intense solar radiation during hot and humid summer periods resulting in higher O3 formation (53). The HGB area is by the side of a port with extensive petrochemical production and refining facilities that emit HRVOCs, resulting in higher O3 formation and accumulation compared with other urban areas having a typical mix of anthropogenic emissions (41, 54). The rate of O3 formation in the HGB area can be as high as 200 ppb per hour in contrast to 40 ppb per hour that is the maximum in other urban areas (55).

Conclusions A review of the O3 air quality studies in HGB area was provided. Accurate determination of the sources of the primary O3 precursors, i.e., NOx and VOCs were found to be essential in accurately predicting the O3 exceedances. One of the principle contributors to high O3 exceedances in the HGB area is the emission of HRVOCs and NOx by petrochemical industrial activities around the Houston Ship Channel. The higher industrial emission of VOCs and NOx coupled with favorable meteorology facilitates the rapid formation and accumulation of O3 concentrations. The use of CAMx is recommended for photochemical modeling of O3 concentrations due to its better performance than CMAQ. The use of CAMx with CB6 chemical mechanism proved better prediction and is recommended for use rather than the use of CAMx with CB5 chemical mechanism.

Acknowledgments This work is supported by the National Science Foundation (NSF) through the Center for Energy and Environmental sustainability (CEES), A CREST grant ( #1036593).

References 1. 2. 3.

Haagen-Smit, A. J. Chemistry and Physiology of Los Angeles Smog. Ind. Eng. Chem 1952, 44, 1342–1346. Leighton, P. A. Photochemistry of Air Pollution; Academic Press: San Diego, CA, 1961; 312 p. Levy, H., II. Normal Atmosphere: Large Radical and Formaldehyde Concentrations Predicted. Science 1971, 173, 141–143. 46 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

4. 5.

6.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

7. 8. 9.

10.

11.

12.

13.

14.

15.

16. 17.

18.

Crutzen, P. J. The Role of NO and NO2 in the Chemistry of the Troposphere and Stratosphere. Ann. Rev. Earth Planet. Sci. 1979, 7, 443–472. Seinfeld, J. H.; Pandis, S. N. Atmospheric Chemistry and Physics: From Air Pollution to Climate Change, 2nd ed.; John Wiley & Sons Inc.: Somerset, NJ, 2006; 1232 p. Amann, M.; Derwent, D.; Forsberg, B.; Hänninen, O.; Hurley, F.; Krzyzanowski, M.; de Leeuw, F.; Liu, S. J.; Mandin, C.; Schneider, J.; Schwarze, P.; Simpson, D. Health Risks of Ozone from Long-Range Transboundary Air Pollution; World Health Organization: Copenhagen, Denmark, 2008; 111 p. Lippmann, M. Environmental Toxicants: Human Exposures and Their Effects, 3rd ed.; John Wiley & Sons Inc.: Somerset, NJ, 2009; pp 317−359 Jerrett, M.; Ito, K.; Burnett, R. Ozone Exposure and Mortality. N. Engl. J. Med. 2009, 360, 2788–2789. Bell, M. L.; Dominici, F.; Samet, J. M. A Meta-Analysis of Time-Series Studies of Ozone and Mortality with Comparison to the National Morbidity, Mortality, and Air Pollution Study. Epidemiology 2005, 16, 436–445. Yang, C.; Yang, H.; Guo, S.; Wang, Z.; Xu, X.; Duan, X.; Kan, H. Alternative Ozone Metrics and Daily Mortality in Suzhou: The China Air Pollution and Health Effects Study (CAPES). Sci. Total Environ. 2012, 426, 83–89. Zhang, Y.; Huang, W.; London, S. J.; Song, G. X.; Chen, G. H.; Jiang, L. L.; Zhao, N. Q.; Chen, B. H.; Kan, H. D. Ozone and Daily Mortality in Shanghai, China. Environ. Health Perspect. 2006, 114, 1227–1232. Ou, C. Q.; Wong, C. M.; Ho, S. Y.; Schooling, M.; Yang, L.; Hedley, A. J.; Lam, T. H. Dietary Habits and the Short-term Effects of Air Pollution on Mortality in Chinese Population in Hong Kong. J. Epidemiol. Community Health 2012, 66, 254–258. Liu, T.; Li, T. T.; Zhang, Y. H.; Xu, Y. J.; Lao, X. Q.; Rutherford, S.; Chue, C.; Luo, Y.; Zhu, Q.; Xu, X. J.; Xie, H. Y.; Liu, Z. R.; Ma, W. J. The ShortTerm Effect of Ambient Ozone on Mortality is Modified by Temperature in Guangzhou, China. Atmos. Environ. 2013, 76, 59–67. Piikki, K.; de Temmerman, L.; Ojanpera, K.; Danielsson, H.; Pleijel, H. The Grain Quality of Spring Wheat (Triticumaestivum L.) in Relation to Elevated Ozone Uptake and Carbon Dioxide Exposure. Eur. J. Agron. 2008, 28, 245–254. Feng, Z. Z.; Kobayashi, K.; Ainsworth, E. A. Impact of Elevated Ozone Concentration on Growth, Physiology, and Yield of Wheat (Triticumaestivum L.): A Meta-Analysis. Glob. Change Biol. 2008, 14, 2696–2708. de Temmerman, L.; Legrand, G.; Vandermeiren, K. Effects of Ozone on Sugar Beet Grown in Open-Top Chambers. Eur. J. Agron. 2007, 26, 1–9. Vandermeiren, K.; Black, C.; Pleijel, H.; de Temmerman, L. Impact of Rising Tropospheric Ozone on Potato: Effects on Photosynthesis, Growth, Productivity and Yield Quality. Plant Cell Environ. 2005, 28, 982–996. Ollerenshaw, J. H.; Lyons, T.; Barnes, J. D. Impacts of Ozone on the Growth and Yield of Field-Grown Winter Oilseed Rape. Environ. Pollut. 1999, 104, 53–59. 47 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

19. de Bock, M.; de Beeck, M. O.; de Temmerman, L.; Guisez, Y.; Ceulemans, R.; Vandermeiren, K. Ozone Dose-Response Relationships for Spring Oilseed Rape and Broccoli. Atmos. Environ. 2011, 45, 1759–1765. 20. Nouchi, I.; Ito, O.; Harazono, Y.; Kobayashi, K. Effects of Chronic Ozone Exposure on Growth, Root Respiration and Nutrient Uptake of Rice Plants. Environ. Pollut. 1991, 74, 149–164. 21. Kobayashi, K.; Okada, M.; Nouchi, I. Effects of Ozone on Dry Matter Partitioning and Yield of Japanese Cultivars of Rice (Oryza sativa L.). Agric., Ecosyst. Environ. 1995, 53, 109–122. 22. Yonekura, T.; Shimada, T.; Miwa, M.; Arzate, A.; Ogawa, K. Impacts of Tropospheric Ozone on Growth and Yield of Rice (Oryza sativa L.). J. Agric. Meteorol. 2005, 60, 1045–1048. 23. Yamaguchi, M.; Inada, H.; Satoh, R.; Hoshino, D.; Nagasawa, A.; Negishi, Y.; Sasaki, H.; Nouchi, I.; Kobayashi, K.; Izuta, T. Effects of Ozone on Growth, Yield and Leaf Gas Exchange Rates of Two Japanese Cultivars of Rice (Oryza sativa L.). J. Agric. Meteorol. 2008, 64, 131–141. 24. Nouchi, I.; Aoki, K. Morning Glory as a Photochemical Oxidant Indicator. Environ. Pollut. 1979, 18, 289–303. 25. Reinert, R. A.; Nelson, P. V. Sensitivity and Growth of Twelve Elatior Begonia-Hiemalis Cultivars to Ozone. Hort. Sci. 1979, 14, 747–748. 26. Bennett, J. H. Foliar Exchange of Gases. In Methodology for the Assessment of Air Pollution Effects on Vegetation: A Handbook; Heck, W. W., Krupa, S. V., Linzon, S. N., Eds.; Air Pollution Control Association: Pittsburgh, PA, 1979; pp 10-1−10-29. 27. Musselman, R. C.; Oshima, R. J.; Gallavan, R. E. Significance of Pollutant Concentration Distribution in the Response of Red Kidney Beans (Phaseolus Vulgaris) to Ozone. J. Am. Soc. Hortic. Sci. 1983, 108, 347–351. 28. Musselman, R. C.; Huerta, A. J.; McCool, P. M.; Oshima, R. J. Response of Beans to Simulated Ambient and Uniform Ozone Distribution with Equal Peak Concentrations. J. Am. Soc. Hortic. Sci. 1986, 111, 470–473. 29. Ashmore, M. R. Effects of Ozone on Vegetation in the United Kingdom. In Proceedings of an International Workshop on the Evaluation and Assessment of the Effects of Photochemical Oxidants on Human Health, Agricultural Crops, Forestry, Materials and Visibility; Grennfelt, P., Ed.; Swedish Environmental Research Institute (IVL): Goteborg, Sweden, 1984; pp 92−104 30. Amiro, B. D.; Gillespie, T. J.; Thurtell, G. W. Injury Response of PhaseolusVulgaris to Ozone Flux Density. Atmos. Environ. 1984, 18, 1207–1215. 31. Tonnaeijck, A. E. G. Effects of Peroxyacetyl Nitrate (PAN) and Ozone on Some Plant Species. In Proceedings of the International Workshop on an Evaluation and Assessment of the Effects of Photochemical Oxidants on Human Health, Agricultural Crops, Forestry, Materials and Visibility; Grennfelt, P., Ed.; Swedish Environmental Research Institute (IVL): Goteborg, Sweden, 1984; pp 118−127. 32. Hogsett, W. E.; Tingey, D. T.; Holman, S. R. A Programmable Exposure Control System for Determination of the Effects of Pollutant Exposure Regimes on Plant Growth. Atmos. Environ. 1985, 19, 1135–1145. 48 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

33. United States Environmental Protection Agency. Review of the National Ambient Air Quality Standards for Ozone; http://yosemite.epa.gov/opei/ RuleGate.nsf/byRIN/2060-AP38#1 (accessed June 3, 2015). 34. United States Environmental Protection Agency. 8-hour Ozone Nonattainment Areas (2008 standard); http://www.epa.gov/oaqps001/ greenbk/map8hr_2008.html (accessed June 3, 2015). 35. Texas Commission on Environmental Quality. Houston-Galveston-Brazoria and the State Implementation Plan; https://www.tceq.texas.gov/airquality/ sip/hgb/sip-hgb (accessed June 3, 2015). 36. Texas Commission on Environmental Quality. Houston-GalvestonBrazoria Ozone SIP Mid-Course Review Modeling: CAMx Modeling Domain; http://www.tceq.state.tx.us/airquality/airmod/data/hgbmcr/hgb1_ camx_domain.html (accessed June 3, 2015). 37. Daum, P. H.; Kleinman, L. I.; Springston, S. R.; Nunnermacker, L. J.; Lee, Y. N.; Weinstein‐Lloyd, J.; Zheng, J.; Berkowitz, C. M. A Comparative Study of O3 Formation in the Houston Urban and Industrial Plumes During the 2000 Texas Air Quality Study. J. Geophys. Res. 2003, 108, D23. 38. Zhou, W.; Cohan, D. S.; Henderson, B. H. Slower Ozone Production in Houston, Texas Following Emission Reductions: Evidence from Texas Air Quality Studies in 2000 and 2006. Atmos. Chem. Phys. 2014, 14, 2777–2788. 39. Murphy, C.; Allen, D. Hydrocarbon Emissions from Industrial Release Events in the Houston-Galveston Area and Their Impact on Ozone Formation. Atmos. Environ. 2005, 39, 3785–3798. 40. Li, G.; Zhang, R.; Fan, J.; Tie, X. Impacts of Biogenic Emissions on Photochemical Ozone Production in Houston, Texas. J. Geophys. Res. 2007, 112, D10309. 41. Ryerson, T. B.; Trainer, M.; Angevine, W. M.; Brock, C. A.; Dissly, R. W.; Fehsenfeld, F. C.; Frost, G. J.; Goldan, P. D.; Holloway, J. S.; Hubler, G.; Jakoubek, R. O.; Kuster, W. C.; Neuman, J. A.; Nicks, Jr., D. K.; Parrish, D. D.; Roberts, J. M.; Sueper, D. T. Effect of Petrochemical Industrial Emissions of Reactive Alkenes and NOx on Tropospheric Ozone Formation in Houston, Texas. Atmos. Environ. 2007, 41, 837–853. 42. Washenfelder, R. A.; Trainer, M.; Frost, G. J.; Ryerson, T. B.; Atlas, E. L.; de Gouw, J. A.; Flocke, F. M.; Fried, A.; Holloway, J. S.; Parrish, D. D.; Peischl, J.; Richter, D.; Schauffler, S. M.; Walega, J. G.; Warneke, C.; Weibring, P.; Zheng, W. Characterization of NOx, SO2, Ethene, and Propene from Industrial Emission Sources in Houston, Texas. J. Geophys. Res. 2010, 115, D16311. 43. Webster, M.; Nam, J.; Kimura, Y.; Jeffries, H.; Vizuete, W.; Allen, D. T. The Effect of Variability in Industrial Emissions on Ozone Formation in Houston, Texas. Atmos. Environ. 2007, 41, 9580–9593. 44. Neuman, J. A.; Nowak, J. B.; Zheng, W.; Flocke, F.; Ryerson, T. B.; Trainer, M.; Holloway, J. S.; Parrish, D. D.; Frost, G. J.; Peischl, J.; Atlas, E. L.; Bahreini, R.; Wollny, A. G.; Fehsenfeld, F. C. Relationship between Photochemical Ozone Production and NOx Oxidation in Houston, Texas. J. Geophys. Res. 2009, 114, D00F08. 49 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch002

45. Daum., P. H.; Kleinman, L. I.; Springston, S. R.; Nunnermacker, L. J.; Lee, Y. N.; Weinstein-Lloyd, J.; Jheng, J.; Berkowitz, C. Origin and Properties of Plumes of High Ozone Observed During Texas 2000 Air Quality Study (TEXAQS 2000). J. Geophys. Res. 2004, 109, D17306. 46. Sarker, S.; Kommalapati, R. R.; Huque, Z. Influence of Reactive Volatile Organic Compounds on Ozone Production in Houston-Galveston-Brazoria Area. J. Environ. Prot. 2015, 6, 399–408. 47. Xiao, X.; Cohan, D. S.; Byun, D. W.; Ngan, F. Highly Nonlinear Ozone Formation in the Houston Region and Implications for Emission Controls. J. Geophys. Res. 2010, 115, D23309. 48. Byun, D. W.; Kim, S. T.; Kim, S. B. Evaluation of Air Quality Models for the Simulation of a High Ozone Episode in the Houston Metropolitan Area. Atmos. Environ. 2007, 41, 837–853. 49. Sarker, S.; Kommalapati, R. R.; Huque, Z. Ozone Simulations for HGB Area Using Different Chemical Mechanisms with CAMx. Int. J. Adv. Res. Sci. Eng. 2015, 4, 458–465. 50. Banta, R. M.; Senff, C. J.; Nielsen-Gammon, J.; Darby, L. S.; Ryerson, T. B.; Alvarez, R. J.; Sandberg, S. P.; Williams, E. J.; Trainer, M. A Bad Air Day in Houston. Bull. Am. Meteorol. Soc. 2005, 86, 657–669. 51. Nielsen-Gammon, J.; Tobin, J.; McNeel, A.; Li, G. A Conceptual Model for Eight-Hour Ozone Exceedances in Houston, Texas Part II: Eight Hour Ozone Exceedances in Houston-Galveston Metropolitan Area; Technical Report for Houston Advanced Research Center; Texas Environmental Research Consortium and the Texas Commission on Environmental Quality: Houston, TX; January 2005. 52. Nielsen-Gammon, J.; Tobin, J.; McNeel, A.; Li, G. A Conceptual Model for Eight-Hour Ozone Exceedances in Houston, Texas Part I: Background Ozone Level in Eastern Texas; Technical Report for Houston Advanced Research Center; Texas Environmental Research Consortium and the Texas Commission on Environmental Quality: Houston, TX; January 2005. 53. Ranpenglück, B.; Perna, R.; Zhong, S.; Morris, G. A. An Analysis of the Vertical Structure of the Atmosphere and the Upper-Level Meteorology and Their Impact on Surface Ozone Levels in Houston, Texas. J. Geophys. Res. 2008, 113, D17315. 54. Kleinman, L. I.; Daum, P. H.; Imre, D.; Lee, Y. N.; Nunnermacker, L. J.; Springston, S. R.; Weinstein-Lloyd, J.; Rudolph, J. Ozone Production Rate and Hydrocarbon Reactivity in 5 Urban Areas: A Cause of High Ozone Concentration in Houston. Geophys. Res. Lett. 2002, 29, 105-1–105-4. 55. Kleinman, L. I.; Daum, P. H.; Imre, D.; Lee, Y. N.; Nunnermacker, L. J.; Springston, S. R.; Weinstein-Lloyd, J.; Rudolph, J. Correction to “Ozone Production Rate and Hydrocarbon Reactivity in 5 Urban Areas: A Cause of High Ozone Concentration in Houston”. Geophys. Res. Lett. 2003, 30, 41–1.

50 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

Chapter 3

Challenges Associated with Using Retorts To Limit Mercury Exposure in Artisanal and Small-Scale Gold Mining: Case Studies from Mozambique, Ecuador, and Guyana Adam M. Kiefer,*,1 Kevin Drace,2 Caryn S. Seney,1 and Marcello M. Veiga3 1Department

of Chemistry, Mercer University, 1400 Coleman Avenue, Macon, Georgia 31207 2Department of Biology, Mercer University, 1400 Coleman Avenue, Macon, Georgia 31207 3Norman B. Keevil Institute of Mining Engineering, University of British Columbia, Vancouver, Canada *E-mail: [email protected].

Artisanal and small-scale gold mining (ASGM) is recognized as the number one source of anthropogenic mercury pollution in the world. Miners use mercury to amalgamate gold, then heat the amalgam to evaporate the mercury. This process, referred to by miners as burning, releases large quantities of mercury vapor that affects the health of miners and community members. A retort is a mercury capture device that provides a simple solution to reduce human exposure to mercury vapor during the burning process. In spite of the low cost, ease of use and numerous outreach programs that have introduced miners to this technology, miners have been reluctant to use retorts. This chapter provides a review of ASGM processes involving mercury, discusses the health effects of mercury vapor, and provides case studies on retort use in Mozambique, Ecuador and Guyana. The chapter concludes with a discussion on reasons why retorts have not been widely adopted in these countries, and it provides potential solutions to address miners concerns with the use of retorts.

© 2015 American Chemical Society In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

Introduction Artisanal and small-scale mining (ASM) is a significant source of employment for men and women in developing nations around the world, with the majority of these miners processing ore to isolate and sell gold (1). The number of people employed in artisanal and small-scale gold mining (ASGM) tracks with the global price of gold, which increased over 400% from 2002-2012 (2). Unfortunately, the transient nature of miners involved in ASGM coupled with the fact that many miners are involved in illegal mining operations makes it difficult to estimate the total number of gold miners throughout the world. From 2002-2011, estimates have ranged from 13-30 million artisanal and small-scale miners globally (3–7), and in 2014 a study by Seccatore et al. estimated that ~16 million artisanal and small-scale gold miners are responsible for 17-20% of global gold production (2). ASM activities, which include ASGM, impact entire regions and communities, not just the miners themselves. In 2011, Hruschka and Echavarria estimated that there were 25 million artisanal miners, and 150-170 million people were indirectly involved in ASM-related activities (6). In 2013, Jønsson and coworkers estimated that there were 9 million people engaged in ASM in Africa, with 54 million people financially dependent on ASM (8). If correct, this estimate implies that ~5% of Africans are dependent on ASM activities for their livelihoods (9). Although gold mining is a viable option for people in developing nations seeking to escape poverty and unemployment, the dramatic increase in ASGM activities has led to numerous challenges to both human and environmental health (7, 10–21). Mining is an inherently dangerous profession for those without appropriate training. The lack of formal engineering training in this sector results in frequent mine collapses (22). In addition, the processing of ore often releases arsenic, cadmium, chromium, mercury and lead into the environment. For example, from 2010-2013, hundreds of Nigerian children died of lead poisoning directly linked to ASGM activities, in spite of the fact that lead was not being actively mined (23, 24). Lead sulfide was present in high concentrations in the ore and released during mining operations, heavily contaminating soil and drinking water in the local community. Perhaps the most notable of the human and environmental health effects are derived from the fact that the vast majority of ASGM workers use elemental mercury to amalgamate gold and silver (25). It is estimated that over 1,600 tonnes of metallic mercury are consumed in ASGM each year (26). The resulting chemical waste is emitted directly into the atmosphere and released into watersheds. ASGM represents 37% of anthropogenic atmospheric mercury emissions globally and is now recognized as the number one source of anthropogenic mercury emissions to the environment (2, 27). Mercury is a bioaccumulative toxin, and in its elemental form mercury is easily distributed throughout the environment. Over time it is oxidized to water-soluble inorganic mercury salts that are then modified by anaerobic bacteria and methylated. These highly toxic organomercury complexes accumulate in fish and are incorporated into higher order predators, including humans (28). For miners, the most immediate danger from elemental mercury during the mining process occurs when mercury is evaporated from the amalgam by heating (11). Miners directly inhale the mercury vapor. Exposure to these mercury vapors 52 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

results in a variety of health effects including tremors, short-term memory loss, kidney damage, and in high concentrations can lead to death (14, 25). A number of outreach programs have provided ASGM workers with retorts, which are a suitable, inexpensive technology that traps up to 95% of mercury during the burning process (8, 29–31). In spite of the low cost and health benefits attributed to retorts, this technology has not been widely adopted by miners (1, 31–36). This chapter will provide the reader with an overview of mercury usage in ASGM activities, a brief overview of programs related to the implementation of retorts in ASGM communities, and three case studies highlighting retort use and misuse in Ecuador, Guyana and Mozambique.

An Overview of ASGM Processing in Developing Nations This section will provide the reader with an understanding of the mining and processing of gold ore in Portovelo and Nambija, Ecuador; Mahdia, Guyana and the Manica Province, Mozambique. Gold mining in Ecuador and Mozambique rely on mechanical comminution (grinding) of hard rock. Mining in Guyana is largely relegated to the mining of colluvium, or geological material that has deposited at the base of a hill due to weathering over millions of years.

Hard Rock Mining: Mozambique In the mining communities of Munhena and Tsetsera in the Manica Province of Mozambique, the majority of gold is mined from hard rock. Miners extract ore along quartz veins in the hills and mountainsides and then process the ore where they, and often their families, reside. While artisanal miners in Mozambique often extract the ore working within small groups, individual miners process their own ore. These miners will initially crush the ore into pieces of approximately 1-2” in diameter using a mortar and pestle. The crushed ore is then transferred into a ball mill, which is a cylinder equipped with stainless steel ball bearings (Figure 1). Often mercury is added to the ball mill at this stage. The mill is turned by hand for approximately one hour, and the ball bearings crush the ore while the mercury amalgamates the liberated gold in a process known as whole-ore amalgamation. Upon completion, the finely-ground contents are transferred to a large plastic basin for panning. The mercury-contaminated gangue is separated and placed on the ground or in small tailings pits, where over time it enters local waterways. Additional mercury is added to the heavier black sands, and after agitation, the mercury is strained through a fine mesh cloth producing a gold amalgam. While the excess mercury is collected and reused, a substantial portion has already been lost to the tailings. The amalgam is often heated or “burned” on a smoldering log. The temperature of the smoldering log is insufficient to rapidly vaporize the mercury; thus, miners resort to blowing on the log until the mercury has evaporated producing sponge gold, or doré (Figure 2). The miners directly inhale evolved mercury into their lungs. Because of the inefficiency of the burning process, the 53 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

sponge gold still contains mercury. The sponge gold can be sold on-site to people that provide the miners with mercury, who then purify the gold in populated areas. The mercury trapped in the sponge gold is released to the environment during this process.

Figure 1. A miner operates a ball mill in Munhena, Mozambique. A ball mill is a cylinder containing stainless steel ball bearings used to crush ore and liberate gold. (see color insert)

Figure 2. A) The amalgam is heated on a smoldering log. B) A miner blows on the log to increase the temperature to heat the amalgam more quickly. During this process he breathes in mercury vapor. (see color insert) 54 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

A few medium-scale mining operations in the region operate in much the same way; however, these operations often have automated jaw crushers and ball mills. Ball mills powered by motors allow for the continuous processing of ore as opposed to the batch processing found in the hand-powered ball mills that artisanal miners use. These ball mills increase throughput and yield, as the automated mills allow for better control of particle size and maximization of gold liberation. Under these conditions the gold is concentrated by panning or centrifugation prior to amalgamation, preventing both whole-ore amalgamation and the contamination of the tailings by mercury. We have previously reported on mercury-free technology being used in the region, but the practice of separating the black sands (magnetite and ball mill iron filings) from gold using magnets is not applicable in most regions of the Manica Province (37).

Hard Rock Mining: Ecuador Similar to Mozambique, in the mining communities of El Oro Province, Ecuador, the majority of gold is associated with sulfides in hard rock. However, the processing of ore rarely occurs at the mine, but rather in the city of Portovelo where more efficient and expensive equipment can be rented to process the ore (7, 38, 39). These processing centers offer miners access to techniques and equipment that they could not afford themselves. Most processing centers offer miners these services for a small fee on the condition that the processing center keeps all of the tailings. Miners extract the ore at their mine, and once a predetermined amount of ore has been collected, the ore is transferred via vehicle to a processing center in Portovelo. Miners select one of the 87 processing centers based on previous experience and the amount of ore that needs to be processed. Artisanal miners with a small amount of rich ore prefer to utilize one or more chanchas to carry out either whole-ore or concentrate amalgamation. Chanchas are small ball mills powered by an electric motor that operate on batch-scale as opposed to continuous flow (Figure 3A). Processing plant owners will often rent chanchas to miners by the hour or batch, with the understanding that the processing plant keeps the tailings. When a sufficient amount of tailings have been accrued, the processing plant owner then processes the tailings with cyanide and recovers the remaining gold. Because whole-ore amalgamation rarely results in more than 30% recovery of gold in ore (30), the processing plant owners often make a large profit from gold-containing tailings. For miners with large quantities of ore, Chilean mills are utilized for comminution (Figure 3B). These mills consist of three heavy cement wheels rimmed with steel that rotate in a circle on a steel plate. After being passed through a jaw crusher, the ore is shoveled into the mill and crushed. Water continuously passes over the material, which is then screened and passed down a sluice angled at ~5° containing nylon carpet. The heavier gold lodges in the carpet, while the lighter materials flow to a tailings pit. The carpet is rinsed approximately every hour into large buckets for amalgamation, and further concentrated by manual panning. Although Chilean mills are continuous throughput, they are still 55 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

inefficient, and a portion of the gold remains in the tailings. However, because only the concentrate is amalgamated the tailings from a Chilean mill can directly undergo cyanidation without the issue of solubilizing mercury as occurs after whole-ore amalgamation.

Figure 3. A) Miners emptying a chancha in Portovelo, Ecuador. B) A Chilean mill in Portovelo, Ecuador. (see color insert)

Independent of the form of comminution, the amalgamation process occurs similarly to the process in Mozambique. The mercury and gold are strained through a tightly woven cloth, and the excess mercury is collected and reused. However, the resulting amalgam is typically significantly larger than amalgams found in Mozambique due to both improved processing practices and higher gold grade of the ore. The resulting spherical amalgam is then crushed flat, mixed with brown sugar and heated with a propane torch to evaporate the mercury. The role of brown sugar is not understood, although miners claim that it increases the rate of evaporation of mercury. Although this process is often performed outside in the open environment or under a makeshift fumehood without proper condensers or filters, occasionally mercury capture devices such as retorts are employed. The resulting doré is then sold to local gold shops in Portovelo and neighboring communities for further refinement. The owners of processing centers leach the collected tailings with cyanide. It has been suggested that cyanidation following whole-ore amalgamation results in the formation of water-soluble cyanomercury species that are more bioavailable than elemental mercury (16). Most final tailings containing these organomercury species and elemental mercury are ultimately disposed of in local streams and rivers. In Portovelo there are approximately 375 cyanidation tanks with capacities ranging from 14-40 m3 (16). In 2011 it was estimated that over 880,000 tonnes/a of tailings containing about 650 kg of Hg and 6000 tonnes of cyanide are discharged into the rivers (40). 56 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

Colluvial Mining: Guyana As rock weathers over a period of millions of years, it travels down hillsides and is deposited as loose, unconsolidated sediment known as colluvium. The natural weathering results in the partial liberation of gold from the surrounding rock, and if the material reaches a stream the material (now alluvium) it is further separated from gangue. In essence, nature grinds the material and separates the gold. The mining community of Mahdia, Guyana, located in the Potaro-Siparuni region has numerous colluvial and alluvial deposits of gold (41, 42). Miners remove the overburden using heavy machinery or by hand, and then use highpressure water flow to loosen the colluvium (Figure 4a). The resulting slurry is then pumped to a carpeted sluice box, where water flow and gravity separate the gold and heavy minerals from the gangue (Figure 4b). The majority of gold is very fine and is not captured on the sluice bed. This practice has led to small-scale miners and “pork knockers” (43), as artisanal miners are referred to in Guyana, reworking the tailings from abandoned claims.

Figure 4. A) Miners created a slurry using high powered water hoses and the slurry is pumped to B) a sluice box. (see color insert)

Miners collect the heavy concentrate from the sluice box in large basins and approximately once a week add mercury to amalgamate the gold. The amalgam is separated from the excess mercury using a fine mesh cloth. The resulting amalgam is burned in open air using a propane torch or is wrapped in wet leaves and heated over a bonfire in close proximity to the miners’ encampment. At some of the medium-scale mines, miners will use a retort to condense and capture the mercury, which is then returned to the original mercury supply. Miners often sell the gold directly to gold shops licensed by the Guyanese government.

Mercury Vapor and Human Health Mercury has been used to amalgamate gold since 1000 BC (44), and mercury usage in ASGM activities remains ubiquitous today. ASGM workers use mercury for numerous reasons (30). First and foremost, mercury works for artisanal and 57 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

small-scale miners. It is both inexpensive and readily available for purchase in mining communities. Moreover, a miner with no formal training can learn to use mercury in ASGM within an hour oftentimes being trained by friends and family members. Although amalgamation is inefficient, and miners more often than not discard the majority of gold with the tailings, many miners can isolate gold from the rest of the ore in a matter of hours using mercury. This transformation translates to food on the table, and a higher quality of life in the short term. Although in some regions, such as in Portovelo, Ecuador, miners have access to mercury-free technologies such as cyanidation and flotation, many miners around the world do not have access to the training needed to operate these systems. In addition, these miners may not have the capital to establish and maintain them. Even when available, many artisanal and small-scale miners cannot collect enough ore to make these processes feasible and economical when they are available. The vast majority of gold miners that use mercury form and burn amalgams, and most do so without taking appropriate safety precautions. It is during the burning process that there is an immediate health risk to the individual miner as well as the mining community as a whole. The effects of chronic mercury exposure associated with gold mining are well documented, and extend not only to miners but their families and members of the surrounding community as well. For example, Bose-O’Reilly and coworkers demonstrated that in the Philippines children “just living” in gold mining areas contaminated by mercury have statistically significant higher levels of mercury in their bodies than the control group (13). Many of the children in this community displayed the symptoms of chronic mercury exposure, particularly ataxia and coordination problems. The authors attribute these symptoms to exposure to mercury vapor as opposed to organomercury species. Table 1 relates mercury concentrations in the air to potential human health effects. It is important to note that during the burning of amalgams, concentrations of mercury can exceed 6,000,000 – 60,000,000 ng/m3 when burned without a retort or at a gold shop (45, 46). Chronic exposure to mercury vapors results in a variety of debilitating impairments over time. Exposure to high concentrations of mercury during the burning process can lead to acute toxicity, to mercurial pneumonitis and to rapid death (51). For example, a manager of a Guyanese gold mine was treated for mercury poisoning after he was exposed to high concentrations of mercury vapor while burning an amalgam (52). Although he was discharged from the hospital after 9 days of chelation therapy with penicillamine, the damage to his lungs caused by acute chemical pneumonitis resulted in his inability to walk one flight of stairs without shortness of breath 11 months after treatment. In 1987, 11 Filipinos became ill and one died after burning a gold amalgam without appropriate ventilation (11). The effects of mercury vapor are not limited to the pulmonary system. The effects of mercury vapor on kidneys of miners are dramatic and cases of kidney failure followed by death are reported in Venezuela (53) and in Colombia, where a high incidence of kidney transplants is observed in towns with high concentrations of Hg-based atmospheric pollution (54). These examples clearly demonstrate that exposure to high concentrations of elemental mercury released during the burning process is a threat to human health in ASGM communities. 58 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Table 1. Potential Symptoms Due to Exposure to Mercury Vapor of Specific Concentrations

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

[Hg0] (ng/m3)

Symptoms

Reference

200

ATSDRa minimal risk level

(47)

25,000

Gingivitis, stomatitis, vision changes, hearing loss, kidney effect, teratogenic effects, tremors, emotional instability and irritability, peripheral neuropathy, reproductive effects (ACGIHb Threshold Limit Value, time weighted average for 8-h work day)

(48, 49)

50,000

Kidney damage, erethism (irritability, weakness, sensitivity to stimulation, shyness, depression, insomnia, and eventually memory loss and tremors), micromercurialism, respiratory effects, chest pains, coughing, dermatitis, damage to lens of eye (NIOSHc Recommended exposure limit, time weighted average for 8-h work day)

(48)

100,000

Mercurial pneumonitis and all of the above (OSHAd Ceiling limit)

(48, 49)

200,000

24-h emergency exposure guidance level (EEGL)

(50)

10,000,000

Immediately dangerous to life

(50)

Agency for Toxic Substances and Disease Registry (ATSDR); American Conference of Governmental Industrial Hygienists(ACGIH); c National Institute for Occupational Safety and Health(NIOSH); d Occupational Safety and Health Administration (OSHA). a

b

Retorts in ASGM There is a simple solution for limiting human exposure to mercury vapor during the amalgam burning process. A retort is a mercury capture device that allows for the safe heating of an amalgam using a propane torch or campfire. As the amalgam is heated, mercury is vaporized, condensed and captured in a secondary container. Retorts must be constructed of materials that will not amalgamate when exposed to mercury vapor; if a retort is constructed of metal, steel must be used (1). Three examples of retorts are shown in Figure 5. Retort A and B are two styles of retort from Mozambique. Retort A consists of a chamber that holds the amalgam that screws into the condenser. It is very similar to the most common “water-pipe” retort designed by Raphael Hypolito (8, 55) but possesses a small, refillable water jacket is welded onto the condenser to ensure more efficient condensation of the mercury vapor. Liquid mercury is captured at the end of the pipe in a secondary container. This style of retort is made locally from discarded pipefittings and costs ~$35 USD. Miners report that it works best with larger amalgams. Retort B is 59 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

referred to as a “kitchen-bowl retort” and is constructed from materials available in many village markets (56–59). The operation of this retort is simple: the amalgam is placed in a stainless steel cup inside a larger salad bowl. If cast iron or steel are used the retorted gold turns brown, but the gold remains yellow when in contact with stainless steel or when the retort is lined with clay (1). This surface brown color should not reduce the gold grade and is easily removed when the gold is hammered, but many gold buyers fool the miners paying less for brown gold. The cup is surrounded by sand in order to facilitate heat transfer. A second inverted bowl is placed on top, and the seam between the two large bowls is sealed using wet sand. The top bowl is initially made of a glass salad bowl that allows visual inspection of the burning process by the miners. The glass bowl can later be replaced by a stainless steel or enameled bowl that cools down faster than the glass when the retorting process ends. The lower bowl is then heated using either a campfire or a propane torch. As the mercury vaporizes, it is condensed on the cooler top bowl and is captured by the sand on the bottom of the bowl. Retort C is one of many designs available in and around Portovelo, Ecuador (60). It is similar in operation to retort A, except it has a built in burner that is directly attached to a propane tank. In addition, the water jacket around the condenser is much larger and allows for a continuous flow of water to ensure more efficient cooling of the mercury vapor.

Figure 5. A) A water-pipe retort in Munhena, Mozambique. B) A miner demonstrates the operation of a kitchen-bowl retort in Munhena, Mozambique. C) A retort in Portovelo, Ecuador with a built-in propane torch. (see color insert)

Retorts such as these have been introduced in many mining communities, and have been demonstrated to capture up to 95% of mercury vapor (8, 30). The kitchen bowl retort can be made for less than $5 USD (57), and yet many miners seem adverse to using them regardless of their low cost, the training they receive and the safety benefits inherent in using retorts. The explanation for this surprising behavior is complex and interwoven through the culture of ASGM. In the next section three independent case studies from Ecuador, Guyana and Mozambique that highlight some of the barriers and prohibitive perceptions to implementing retorts in ASGM communities will be examined. 60 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

Case Study #1: Mozambique In 2005, a team from the University of British Columbia (UBC) developed and implemented an educational pilot-program for miners in the Munhena region of the Manica province in Mozambique (57–59). This program was specifically designed to instruct miners how to use retorts when burning amalgams. Prior to intervention, amalgams were burned in open air in bonfires. This program was organized and delivered by experts and was supported by the local mining association. At the time of the educational session, the mining association represented 3,764 miners in the Munhena region, and it was estimated that over 12,000 people in the surrounding area were dependent on ASGM activities. Mining in Munhena was exclusively artisanal in 2005. Ore was extracted at the top of the hill in open pit mines angled into the hillside. The ore was processed in the lowest part of the valley where there were year-round streams that aided in processing. Miners exclusively amalgamated the concentrate after comminution and panning. Between the mines and the processing area was a village where some miners and their families lived. The village was comprised of ~25% women, some of whom processed ore but never collected ore from the mine. Children under 15 years of age were also seen processing ore, although they too were excluded from collecting ore in the mines. Because the mines were open, unreinforced pits dug into the hillside, death from collapses and landslides were common. Due to the dangers associated with the mine, miners started reprocessing tailings and waste rock to glean remaining gold from processing as opposed to mining new ore. Amalgams formed during processing were burned both in the village and at the processing site. Air concentrations of mercury were so high during the initial visit (>50,000 ng/m3) that the Lumex spectrometer used to measure mercury became contaminated and required repair work and recalibration. For these reasons, the UBC team developed an educational experience that included three workshops highlighting chemical hygiene and the appropriate use of retorts. These workshops were attended by a select group of miners, who were then tasked with training the community as a whole. This model, referred to as “training-the-trainer”, engages members of the community in active participation in their own education and provides miners with a sense of purpose and inclusion. The first workshop in 2005 introduced miners to the hazards of working with mercury. Miners were taught about the dangers of inhaling mercury vapor during the burning process, and an informal evaluation of exhaled air using the Lumex spectrometer demonstrated to miners that they were inhaling tremendous amounts of mercury during processing. Of 25 miners sampled, the average miner’s exhaled breath had a concentration of 8,200 ng/m3 (57, 59). Three miners that identified they recently burned amalgams had breath concentrations of 50,000 ng/m3, 60,000 ng/m3 and 27,000 ng/m3. The miner with a breath concentration of 27,000 ng/m3 was identified by the team as a child of less than 15 years of age. It is important to note that although the best way to analyze human exposure to elemental mercury is through urinalysis (12, 13, 61), breath analysis has been demonstrated to provide a reasonable approximation of mercury exposure (62). Given the remoteness of the mining village, breath analysis was used exclusively for demonstration purposes to show miners in real-time their exposure to mercury vapor. 61 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

The second workshop introduced miners to both the kitchen-bowl retort and water-pipe retort. This group of miners was taught the underlying principles of how retorts work, how to safely use retorts and the advantages of using retorts properly. The retorts were demonstrated, and the surrounding air analyzed for mercury content. The kitchen-bowl retort was particularly well accepted, because miners could watch the amalgam as mercury was driven off and the gold revealed. During eight trials of the kitchen bowl retort, it was demonstrated that air concentrations two meters away from the retort (the approximate distance from a miner’s face to the retort) never exceeded 1,000 ng/m3. The third workshop introduced miners to mercury-free magnetic sluice boxes and other mercury-free technologies. In addition, miners were taught how to build retorts from locally sourced materials. Upon completing these three workshops in 2005, the miners then relayed their knowledge to other miners in the community under the observation of the UBC team. In spite of the success of the program, the UBC team unequivocally noted that future fieldwork was needed to assess the miners’ health and that continued monitoring of the miners’ usage of retorts was necessary to ensure that safe and appropriate techniques were being used. The team also suggested isolating one area away from the village to burn amalgams (59). Due to lack of funding, follow-up on the implementation of retorts did not occur until 2010-2011 when a team from Mercer University evaluated the mine at Munhena. Since the initial report of mining activities at the Munhena mine in 2005, many changes occurred within the mining community that led to dramatic reductions of mercury in the ambient air. This reduction in mercury occurred for two reasons: 1) the splitting of the mine into two distinct entities and 2) the utilization of locally made retorts. Shortly after discovery of gold at Munhena, an association of miners was formed to monitor gold production on site. When the pilot-project occurred in 2005, little had changed at Munhena, with the majority of gold being collected by ~3,700 independent and largely unregulated miners. During the pilot-project, it was noted that miners were forced to reprocess tailings and waste rock because the mine itself was unstable. In many ways, the mine itself was inoperable because new ore could not be safely collected. Shortly after the pilot-project in Munhena, the local mining association sought external investment in the mine (63, 64). The formalization of mining at Munhena resulted in two distinct mining zones, referred to as “Upper Munhena” and “Lower Munhena.” Upper Munhena was converted into a highly organized mining operation with 108 salaried employees. Each was assigned a specified job at the mine. The majority of employees were responsible for extracting ore. In order to improve both the efficiency and safety of the mining process, it was necessary to tunnel into the hill. The miners were provided with appropriate safety equipment, and the shafts were ventilated. The mine also purchased a jaw crusher and a continuousflow ball mill (Figure 6A). The gold was separated on a sluice box, the concentrate was amalgamated, and the tailings were collected in tailings ponds. In 2011, a cyanidation plant was being constructed on site, but progress was slowed due to the difficulty of purchasing the equipment and having it delivered to the site. 62 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

Figure 6. A) Miners crush ore at Upper Munhena in a jaw crusher. A ball mill is located to the right of the miners. B) A miner in Lower Munhena reprocesses tailings. (see color insert)

The Upper Munhena mine employed retorts to capture mercury vapor during the burning process. Amalgams were burned at one location, once a week, as had been suggested in 2005. Only two employees were allowed to burn the amalgams. Both employees were able to demonstrate the appropriate use of the water-pipe retort, which was of the same design introduced in 2005. A cursory inspection of the retort revealed that it was often used. When the lower chamber was removed, the concentration of mercury 1 m away exceeded 20,000 ng/m3. Although the two employees were able to explain the operation of the kitchen bowl retort, it was unused at Upper Munhena because the amalgams were large and more easily burned using the pipe retort. In 2010 and 2011 at Upper Munhena, the concentration of mercury detected never exceeded 200 ng/m3 except for the burning area, which never exceeded 5,000 ng/m3. Shandro and coworkers noted that ambient concentrations in Munhena in 2005 were recorded as high as 30,000 ng/m3 (58). Lower Munhena remains largely unchanged from 2005. The organization of Upper Munhena led to a dramatic decrease in both mercury emissions and exposure of miners to mercury but it also led to a dramatic decrease in the number of miners in Lower Munhena. At a peak of ~3,700 miners at Munhena in 2005, there are now fewer than 100 miners in Lower Munhena. Lower Munhena operates independently of Upper Munhena, and miners still process tailings and waste rock using communal hand-operated ball mills (Figure 6B). Of the miners at Lower Munhena and Upper Munhena in 2010 and 2011, none had been present at the original pilot-project. A representative of the local mining association stated that many miners left Lower Munhena and had moved on to other mining sites in the region. 63 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

From 2005-2010, gold processing at Lower Munhena changed greatly. In 2005 there was virtually no whole-ore amalgamation at Munhena, but in 2011 every single hand-operated ball mill registered mercury concentrations greater than 20,000 ng/m3, indicating the use of mercury during comminution. This process, referred to as whole-ore amalgamation, is environmentally destructive and unnecessary, as mercury is pulverized and lost to the tailings (65). When questioned about why miners were performing whole-ore amalgamation, miners responded that this practice was common in Zimbabwe (66). Because of the continual turnover in ASGM communities and political turmoil in Zimbabwe, there has been an increase in the number of unregistered Zimbabwean miners in the Manica Province. Zimbabwe borders Mozambique, and a major border crossing is less than 20 km away from the city of Manica. Zimbabweans and Mozambicans on both sides of the border share a common language and culture. Lower Munhena had a designated area for burning amalgams, but miners continued to burn amalgams in public and residential areas. One such miner proudly showed us his domicile constructed of wood and mud, where he burned his amalgams on the floor over an open fire. Upon being asked why he chose to burn his amalgams in his hut, he stated he did so due to security concerns. The concentration of mercury exceeded 50,000 ng/m3 of mercury outside of his house after burning an amalgam. In 2011 an information-gathering session in Lower Munhena consisting of ~30 miners was conducted. The majority of miners were unfamiliar with the use of retorts. Those miners who were familiar with them chose not to use them. Miners reported that: 1) gold was lost during the burning of amalgams using retorts; 2) the gold turned brown when using retorts; and 3) their amalgams were not large enough to warrant the use of retorts. Miners were informed that gold was not lost during the burning process, but this remains a common misconception in ASGM communities. Gold can turn brown when inappropriate metals are used for the construction of a retort. Upon further inquiry, the miner stating this issue had not experienced the color change himself but heard about it from another miner. Finally, miners were convinced that small amalgams contained insufficient mercury to warrant the use of retorts. A few miners who were familiar with the use of retorts stated that no mercury was captured when the retort was used. In 2014, we published a paper on mercury content in amalgams purchased directly from miners at Munhena and demonstrated that the amalgams contained between 40-60% mercury by mass (67). This number translated to less than 0.63 g of mercury evolved from the most mercury-rich amalgam. It is unlikely that miners would adequately recover this amount of mercury using a retort of any design. However, lack of recovery of mercury does not equate to lack of exposure to mercury vapor. The local miners trained in 2005 have not transferred their knowledge to other mining sites in the Manica Province. Approximately 50 km away from Munhena, Tsetsera is a mining community where miners are largely unaware of the benefits of using retorts. Most miners in Tsetsera refer to mercury as “silver”, perform whole-ore amalgamation and never received formal training of any kind in mining. All processing activities occurred in the village itself, and amalgams were burned throughout the village. Because of a lack of education on the dangers 64 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

of mercury, miners stored their elemental mercury inside their living quarters. Even huts where the inhabitants clearly stated they did not burn amalgams recorded mercury concentrations between 5,000 and 100,000 ng/m3. These huts were mostly stick and mud buildings with thatched roofs. An interesting phenomenon arose in measuring mercury both inside and outside of these huts. Mercury concentrations increased dramatically when the Lumex analyzer was moved from the floor to the roof of the hut. Mercury vapor condenses and adsorbs onto a variety of materials, and it was initially assumed that the increase in concentration was due to the normal evaporation and condensation of mercury during burning. Upon closer inspection, however, the thatched roof seemed to trap mercury more efficiently than the smooth walls. It is assumed that because the thatch has significantly more surface area than the walls, the thatch acts like theoretical plates in a distilling column, allowing for increased condensation of elemental mercury. If this is the case, it follows that the problems of burning amalgams in a hut are compounded by the design and construction of the hut itself. This may lead to increased concentrations of mercury in the air due to inappropriate storage of elemental mercury. Case Study #2: Ecuador Although retorts were introduced during training sessions in the Portovelo region when the Government of Switzerland initiated a Swiss Technical Cooperation (COTESU) project in 1993 (68), many miners continue to burn amalgams in the open air. The district of El Pache in Portovelo has been studied extensively due to the high concentration of mercury vapor in the air (39); however, as part of an air quality assessment from 2013-2015, the concentration of mercury on the main street that runs through the district was determined to rarely exceed 200 ng/m3. For security reasons, the processing plants have built large walls separating themselves from the main street that blocked the flow of air from the processing plants to the street. However, when miners or processing plant workers burn amalgams without using a retort, the concentration of mercury on the street rapidly exceeds 50,000 ng/m3. Within one particular processing plant, the levels of mercury concentration 12-18 h after the last amalgam was burned remained in excess of 100,000 ng/m3, and exceeded 1,000,000 ng/m3 10 m away from the burning area when an amalgam was being heated. Ecuadorian miners routinely state two reasons why they choose not to use retorts. First off, miners do not like losing sight of their gold during the burning process. Because miners’ weekly or monthly salaries are often exclusively based off of the amount of gold they are able to collect, the thought of losing any gold to the retort is terrifying. Second, miners also routinely invoke loss of time as a major factor for why they burn amalgams without using a retort. In order to use a retort effectively, the amalgam must be heated for a set amount of time to ensure all of the mercury is liberated from the sponge gold. Depending on the design, this may exceed an hour. The retort must then be allowed to cool prior to opening the chamber. If the chamber is not cool, copious amounts of mercury will be released and inhaled by the miner when he/she opens the warm retort. Burning an amalgam in the open air using a propane torch often takes less than 15 min. A 30-g amalgam 65 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

in a kitchen-bowl retort with steel covering bowl and thin stainless steel cup can be retorted and cooled down in 30 min, but this is still a long time for an anxious miner. In addition to the time required to burn the amalgam, mercury condensed during the retorting process is less effective at amalgamating gold due to oxidation and the fact that it condenses as small droplets and loses its coalescence. The mercury can be reactivated through an electrolytic process, but this process is also time consuming (1). In spite of outreach and educational programs teaching miners why retorting is important, miners continue to choose convenience over safety. Urban centers such as Portovelo are prime locations for educational outreach on mining safety because miners routinely come to the region in order to process their ore. However, there remain numerous small, hard rock mining settlements throughout Ecuador that continue to process ore at the mine site, which is often located in close proximity to their encampment. These sites are routinely overlooked for training because of their remoteness and lack of easy access. The mining community of Nambija, located in the Zamora-Chinchipe province of Southeastern Ecuador, is one such community. Although mining in the region had occurred for over 500 years, a major gold strike in Nambija caused the town to grow from a population of 100 to 8,000 miners seemingly overnight (69). In spite of the fact that this was a major ASGM site in Ecuador, its illegal status and inaccessibility coupled with excessive violence in the community prevented the effective training of miners in safety protocols. Partial and incomplete training is often more detrimental than no education. In 1995 the Ecuadorean Institute of Mining designed a pilot program to train miners in Nambija to use retorts (68). The program was abandoned prior to completion, and in 2003 Ramirez Requelme and coworkers reported that retorts were either sparsely used or not used at all in Nambija (70). In 2013, a research team from Mercer University traveled to Nambija and were surprised when a miner offered to demonstrate the use of a retort of the miner’s own design (Figure 7). The miner placed his amalgam on a raised iron platform in the middle of a steel bowl. Water was then added to the steel bowl until the water level was only ~1 cm from the top of the platform. A tuna can was then heated for two min using a propane torch and allowed to cool. The miner explained that this was to remove any remnants of plastic sealer from the inside of the can, which caused the gold to be discolored if it wasn’t removed. The can was then floated over the amalgam and heated with the propane torch. In less than 1 min, the air concentration of mercury went from ~200 ng/m3 to 20,000 ng/m3 as measured by a Lumex RA-915M mercury analyzer. After one min, the concentration exceeded the 50,000 ng/m3 maximum detection limit of the instrument, and the Lumex was removed from the area in order to prevent damage to the detector. The instrument was still registering over 50,000 ng/m3 over 9 m from the burning site. It is clear that this modified retort was ineffective. The design is similar to what is referred to as a “tin-fish-tin” retort design (31, 45). However, the tin-fish-tin retort is heated from the bottom, and the fish-tin top is not floated on water. In the case of the Ecuadorian design, when heat is applied to the top of the tuna can that is floated on the water, it breaks the seal between the upper container and the water. This releases a large amount of mercury vapor into the environment. The miner who demonstrated the use of this retort explained that he could complete 66 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

the process in less than 10 min. When the water was removed by the miner, there was elemental mercury that had indeed condensed and dropped to the bottom of the large steel bowl. The miner was informed that although he was able to collect mercury evolved during the burning process, his retort was ineffective at preventing him from breathing in mercury. The miner simply smiled, grabbed a large pan, scooped soil from the ground of the processing area and proceeded to pan the soil. In less than one minute the miner revealed that the dirt alone held multiple grams of elemental mercury.

Figure 7. A) The amalgam is placed on a raised platform surrounded by water. B) The amalgam is covered with a tuna can and heated using a propane torch. C) The sponge gold after heating. (see color insert)

Case Study #3: Guyana Guyana is a small, independent South American country with vast mineral resources and a long history of mercury use in ASGM activities (33, 41, 71–75). The interior of the country is difficult to access and sparsely populated. The PotaroSiparuni region (region 8) has an area of 20,000 km2, with a population in 2012 of only 10,000 residents. The region has numerous, independent small-scale colluvial mining operations that rely on amalgamation to concentrate the gold. The usage of retorts is mandated by law in Guyana, but enforcement of this law is not feasible considering the huge geographic area and sparse population. In 2015 seven mining sites were visited and monitored for airborne mercury concentrations using a Lumex RA-915M AAS. Five of these sites were classified as “small-scale” by the Guyanese government, and the remaining two were classified as “mediumscale”. Artisanal mining is not a recognized term in Guyana; however, small-scale operations in Guyana are comparable to artisanal mining in other nations. All seven sites used land dredges and sluice boxes to concentrate their gold and then amalgamated the concentrate. All mining camps were located within 100 m of mining operations, with miners living in canvas tents on site. Of the seven mining sites evaluated in Potaro-Siparuni, five operations were able to present retorts for inspection, and one mining operation did not possess a retort. The seventh site was not fully operational at the time of the visit, and the mine operators were not present. The largest operation visited clearly used the retort on site. The air concentration of mercury in the encampment never exceeded 67 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

the 1,000 ng/m3, except when the manager opened his mercury storage area and the retort for inspection. Within 3-m of the opened retort, the air concentrations exceeded 20,000 ng/m3, indicating that the retort was recently used. The other medium-scale operation also had a dedicated area for burning amalgams away from living quarters; however, this operation burned amalgams in a stainless steel box with a chimney. This setup is also common in processing plants and gold shops in Ecuador and relies on the heat of the propane torch to carry the mercury vapor through the chimney and away from the face of the miner doing the burning. While this does partly protect the miner burning the amalgam, the mercury vapor is distributed by the wind throughout the camp and is not captured. In addition, the chimney had been moved under a canvas tent to keep it out of the rain. Although the encampment self-reported they burned no more than once a week under the open air, the tent containing the chimney exceeded 20,000 ng/m3 3 m away from the opening. Of the three remaining small-scale sites with retorts, all miners insisted that retorts were used when burning amalgams. However, all living quarters and food preparation areas registered greater than 5,000 ng/m3. It was common for the miners to open the chamber of the retort and the Lumex to register lower readings than that of the rest of the tent, indicating that the retort had not been used. Physical inspection of the retorts showed no heating marks on the outside. All of these sites used the same style and make of retort, similar to retort A (Figure 5) but without a cooling jacket. The condensation tube is very short, and miners had been trained to dip the open end of the retort in a bucket of water to aid in condensation. When pushed for a response as to why the retorts were not used in the field, most miners continued to insist that they were. However, two miners provided the following explanations. First and foremost, the miners stated that the chamber of the retort was too large for the amalgams their operation produced. When used, they never recovered mercury, and the short condenser meant that plastic water buckets intended to trap the mercury melted near the heat source. Even after heating for over an hour, the miners were never able to remove the mercury entirely and had to reheat the amalgam to drive off the remaining mercury. One miner said that it was faster and more efficient to heat the amalgam over a campfire, and it produced better results. The manager at the site without a retort stated that he wrapped his amalgam in wet leaves and burned the amalgam in a campfire directly outside of his living area. Not surprisingly, the campfire area exceeded 10,000 ng/m3, but the fire itself was downwind from the camp’s living area, which registered less than 500 ng/m3.

Conclusion In spite of the fact that retorts can be used to prevent miners’ exposure to elemental mercury, they have repeatedly been dismissed by miners in ASGM communities around the world. The case studies presented herein highlight many of the reasons why. Table 2 highlights the problems associated with retorts in Ecuador, Guyana and Mozambique and provides potential solutions to these issues. 68 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Table 2. Potential and Perceived Problems with the Use of Retorts and Potential Solutions Country

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

Problems or perceived problems associated with retorts and their implementation

Potential Solution

Mercury not recovered, miners do not recognize that small amounts of mercury are dangerous

Mozambique, Guyana

Use a retort designed for the burning of small amalgams, use a retort with a cooling jacket, appropriate educational intervention and follow-up.

Not all mercury is eliminated when using a retort

Guyana

Use improved heating techniques

It takes too long to use a retort

Guyana, Ecuador, Mozambique

Use improved heating techniques, select a more efficient retort, ensure retorts are constructed properly

Miners do not like to lose sight of gold when burning amalgam

Mozambique, Guyana, Ecuador

Kitchen-bowl retort with clear glass cover allows miners to monitor burning

Homemade modifications made to retorts to improve time needed for heating

Ecuador

Follow-up visits to sites to ensure appropriate research use

Relatively high cost for a retort vs. safety

Mozambique

Kitchen-bowl retort

Influx of new and untrained miners

Mozambique

Local, permanent training centers

Security concerns, burning amalgams in public

Mozambique

Develop burning centers in mining communities away from populated areas

Gold is lost during the burning process

Mozambique, Ecuador

This is a fallacy spread by miners when they cannot see the burning process inside the retort. Education and kitchen-bowl retorts. Continued on next page.

69 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Table 2. (Continued). Potential and Perceived Problems with the Use of Retorts and Potential Solutions Country

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

Problems or perceived problems associated with retorts and their implementation

Potential Solution

Gold turns brown in retort

Mozambique

Use stainless steel cups or make a bed of clay inside the retort

Condensed mercury loses coalescence which decreases its ability to form an amalgam

Ecuador

Re-activate the mercury using electrolytic process

Getting miners to use retorts begins with appropriate design and construction. A retort that is not designed for the size of the amalgams being heated will invariably lead to problems. A retort that does not have an appropriate condensation tube will prevent miners from using it appropriately. Training is essential. In spite of their simplicity, using a retort incorrectly is often as dangerous as not using a retort at all. Appropriate technique is essential for successfully volatilizing the mercury and condensing the vapor. For example, miners in Guyana and Mozambique complained that using a retort took too long over a campfire. A prepared trainer can address this problem during training by working with miners to increase the temperature of the fire used to heat the amalgam. In Zimbabwe, miners developed inexpensive bellows (referred to as a mvuto) constructed from a steel pipe and bag that reduced the burning time considerably simply by increasing airflow to the coals (1). Although this is a simple solution, if this concept is not taught to miners during their initial training, miners will choose to discard the retort as opposed to solving the problem. It is also important to distinguish between training and practice. Miners must practice using the retort in front of trainers in order to address potential future issues and to ensure bad habits and “shortcuts” are not developed. Miners are remarkably efficient and results based. It has been noted that miners do not “trouble shoot” problems with retorts, they merely discard them. It may be that one bad experience with a retort is enough to dissuade a miner from using a retort in the future. As was noted in Mozambique, miners openly speak of their experiences using technology. In areas with limited to no formal training opportunities miners may very well dismiss retorts altogether. In areas where educational sessions have occurred, appropriate follow-up is necessary. Educators must return to the site repeatedly to monitor retort usage. Oftentimes this is not financially feasible. The solution may be to construct 70 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

regional training centers in a centralized location so that miners can continuously have access to educational opportunities from a variety of sources, both local and external (76). In addition, miners trained during pilot projects can use the training center as a base of operations for future classes and follow-up. These training centers may also provide meeting sites to improve relationships between miners and governmental officials, allowing for governmental employees to be trained as instructors for future educational development and monitoring ongoing educational initiatives. In some cases, these initiatives do not require external initiation. The Guyana Geology and Mines Commission (GGMC) and the Guyana Mining School and Training Center (GMSTC) are currently working to develop a training center and curricula to train miners on improving gold yield while minimizing mercury use and pollution. While interventions related to retorts in ASGM communities around the world have been largely ineffective to date, retorts are inexpensive, simple to construct and easy to use. Retorts have the potential to keep miners and mining communities safe from mercury vapor exposure. Seemingly, any disadvantages of using retorts in ASGM pale in comparison to the benefits and yet miners continue to refuse to adopt them. Jønsson and coworkers recently commented that “…while miners’ reluctance may partly have to do with inadequate education and an ignorance stemming from a life with strong adversities, their attitude towards the retort, despite its seemingly numerous advantages, needs to be taken seriously (8).” The education process must reach the miners, processors and community members in their environment, using their language and culture. The United Nations Industrial Development Organization (UNIDO) Global Mercury Project in Zimbabwe addressed the importance of retorts while entertaining 8800 people in mining communities using street theater (77). The play Romeo and Juliet was adapted for the African mining community, highlighting the dangers of mercury and the importance of using retorts. The plot involved the son of a successful gold dealer and the daughter of a farmer whose family farm was overrun by illegal miners. When the two lovers admit their love for one another, the resulting discord between the families results in the young woman moving in with the young man. Over time, they both develop mercury intoxication and debilitating symptoms. Regretting the decisions that drove their children away, the farmer blesses the marriage and the gold dealer adopts technology to provide a safer environment for the community. After the play, retorts and efficient gold recovery sluice carpets were demonstrated for the crowd. Unfortunately, the effectiveness of this approach was not validated due to political instability in Zimbabwe; however, the number of people reached was impressive and future work in this area may lead to an effective and reproducible method of delivering training. In spite of their effectiveness, there is a dearth of reports in the peer-reviewed literature on approaches to training miners on how and why they should use retorts. A coordinated effort must be made to develop a connected community of trainers that can share both positive and negative experiences. Until the trainers work together to develop a strategy to address miners’ reluctance to implement retorts, little progress can be made in addressing this issue.

71 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Acknowledgments

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

AMK, KD and CSS acknowledge generous funding through the Mercer on Mission program at Mercer University, and the hard work of 34 undergraduate researchers in the collection of data in Mozambique and Ecuador. AMK, KD and MV are thankful for generous financial support from UNIDO (Project No. 100271, Ecuador) and the Inter-American Development Bank (Guyana Project GY-T1106). AMK thanks Ms. Diane McDonald (Head, Mineral Processing Unit, Guyana Geology and Mines Commission) for helpful conversations regarding ASGM activities in Guyana.

References 1.

Veiga, M. M.; Angeloci-Santos, G.; Meech, J. A. Review of barriers to reduce mercury use in artisanal gold mining. Extr. Ind. Soc. 2014, 1, 351–361. 2. Seccatore, J.; Veiga, M.; Origliasso, C.; Marin, T.; De Tomi, G. An estimation of the artisanal small-scale production of gold in the world. Sci. Total Environ. 2014, 496, 662–667. 3. Quiroga, E. R. The case of artisanal mining in Bolivia: Local participatory development and mining investment opportunities. Nat. Resour. Forum 2002, 26, 127–139. 4. Hentschel, T.; Hruschka, F.; Priester, M. Artisanal and Small-scale Mining: Challenges and Opportunities; International Institute for Environment and Development: London, 2004. 5. Hinton, J. J.; Veiga, M. M.; Beinhoff, C. Women mercury and artisanal gold mining : Risk communication and mitigation. J. Phys. IV 2003, 107. 6. Hruschka, F.; Echavarria, C. Rock Solid Chances for responsible artisanal mining; Series on Responsible ASM; Alliance for Responsible Mining: Medellin, Columbia, 2011. 7. Veiga, M. M.; Angeloci, G.; Hitch, M.; Colon Velasquez-Lopez, P. Processing centres in artisanal gold mining. J. Clean. Prod. 2014, 64, 535–544. 8. Jønsson, J. B.; Charles, E.; Kalvig, P. Toxic mercury versus appropriate technology: Artisanal gold miners’ retort aversion. Resour. Policy 2013, 38, 60–67. 9. Haub, C.; Kaneda, T. 2013 World Population Data Sheet; Population Reference Bureau: Washington, DC, 2013. 10. Grandjean, P.; White, R. F.; Nielsen, A.; Cleary, D.; de Oliveira Santos, E. C. Methylmercury neurotoxicity in Amazonian children downstream from gold mining. Environ. Health Perspect. 1999, 107, 587–591. 11. Eisler, R. Health risks of gold miners: a synoptic review. Environ. Geochem. Health 2003, 25, 325–345. 12. Yard, E. E.; Horton, J.; Schier, J. G.; Caldwell, K.; Sanchez, C.; Lewis, L.; Gastaňaga, C. Mercury exposure among artisanal gold miners in Madre de Dios, Peru: a cross-sectional study. J. Med. Toxicol. 2012, 8, 441–448. 72 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

13. Bose-O’Reilly, S.; Lettmeier, B.; Matteucci Gothe, R.; Beinhoff, C.; Siebert, U.; Drasch, G. Mercury as a serious health hazard for children in gold mining areas. Environ. Res. 2008, 107, 89–97. 14. Risher, J.; World Health Organization; United Nations Environment Programme; International Labour Organisation; Inter-Organization Programme for the Sound Management of Chemicals; International Program on Chemical Safety. Elemental mercury and inorganic mercury compounds: human health aspects; World Health Organization: Geneva, 2003. 15. Betancourt, O.; Narváez, A.; Roulet, M. Small-scale Gold Mining in the Puyango River Basin,Southern Ecuador: A Study of Environmental Impacts and Human Exposures. EcoHealth 2005, 2, 323–332. 16. Velásquez-López, P. C.; Veiga, M. M.; Klein, B.; Shandro, J. A.; Hall, K. Cyanidation of mercury-rich tailings in artisanal and small-scale gold mining: identifying strategies to manage environmental risks in Southern Ecuador. J. Clean. Prod. 2011, 19, 1125–1133. 17. Appleton, J. D.; Williams, T. M.; Orbea, H.; Carrasco, M. Fluvial Contamination Associated with Artisanal Gold Mining in the Ponce Enríquez, Portovelo-Zaruma and Nambija Areas, Ecuador. Water, Air, Soil Pollut. 2001, 131, 19–39. 18. Feng, X.; Dai, Q.; Qiu, G.; Li, G.; He, L.; Wang, D. Gold mining related mercury contamination in Tongguan, Shaanxi Province, PR China. Appl. Geochem. 2006, 21, 1955–1968. 19. Richard, M.; Moher, P.; Telmer, K. Health Issues in Artisanal and Small-Scale Gold Mining: Training for health professionals, (Version 1.0); Artisanal Gold Council: Victoria, BC, 2014. 20. Hylander, L. D. Gold and amalgams: Environmental pollution and health effects; Elsevier: Burlington, 2011. 21. Kristensen, A. K. B.; Thomsen, J. F.; Mikkelsen, S. A review of mercury exposure among artisanal small-scale gold miners in developing countries. Int. Arch. Occup. Environ. Health 2014, 87, 579–590. 22. Basu, N.; Clarke, E.; Green, A.; Calys-Tagoe, B.; Chan, L.; Dzodzomenyo, M.; Fobil, J.; Long, R. N.; Neitzel, R. L.; Obiri, S.; Odei, E.; Ovadje, L.; Quansah, R.; Rajaee, M.; Wilson, M. Integrated Assessment of Artisanal and Small-Scale Gold Mining in Ghana—Part 1: Human Health Review. Int. J. Environ. Res. Public. Health 2015, 12, 5143–5176. 23. Goldman, L.; Pendergrass, J.; Roche, D.; Subramanian, N.; Amsalem, J. Artisanal and Small-Scale Gold Mining in Nigeria: Recommendations to Address Mercury and Lead Exposure; ELI Project No. 121001; Environmental Law Institute: Washington, DC, 2014. 24. Dooyema, C. A.; Neri, A.; Yi-Chun, Lo; Durant, J.; Dargan, P. I.; Swarthout, T.; Biya, O.; Gidado, S. O.; Haladu, S.; Sani-Gwarzo, N.; Nguku, P. M.; Akpan, H.; Idris, S.; Bashir, A. M.; Brown, M. J. Outbreak of Fatal Childhood Lead Poisoning Related to Artisanal Gold Mining in Northwestern Nigeria, 2010. Environ. Health Perspect. 2012, 120, 601–607.

73 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

25. Gibb, H.; O’Leary, K. G. Mercury Exposure and Health Impacts among Individuals in the Artisanal and Small-Scale Gold Mining Community: A Comprehensive Review. Environ. Health Perspect. 2014. 26. Koekkoek, B. Measuring global progress towards a transition away from mercury use in artisanal and small-scale gold mining. Master of Arts in Environment and Management, Royal Roads University: Victoria, BC, Canada. 27. Global Mercury Assessment 2013: Sources, Emissions, Releases and Environmental Transport; UNEP Chemicals Branch, United Nations Pubns: Geneva, Switzerland, 2013. 28. Mercury Fate and Transport in the Global Atmosphere; Mason, R., Pirrone, N., Eds.; Springer U.S.: Boston, MA, 2009. 29. Richard, M.; Moher, P.; Rossin, R.; Telmer, K. Using Retorts to Reduce Mercury Use, Emissions and Exposures in Artisanal and Small Scale Gold Mining: A Practical Guide, (version 1.0); Artisanal Gold Council: Victoria, BC, 2014. 30. Telmer, K.; Stapper, D. Reducing Mercury Use in Artisanal and Small-scale Gold Mining: A Practical Guide; United Nations Industrial Programme (UNEP), 2012. 31. Hinton, J. J.; Veiga, M. M.; Veiga, A. T. C. Clean artisanal gold mining: a utopian approach? J. Clean. Prod. 2003, 11, 99–115. 32. Hilson, G.; Pardie, S. Mercury: An agent of poverty in Ghana’s small-scale gold-mining sector? Resour. Policy 2006, 31, 106–116. 33. Hilson, G.; Vieira, R. Challenges with minimising mercury pollution in the small-scale gold mining sector: Experiences from the Guianas. Int. J. Environ. Health Res. 2007, 17, 429–441. 34. Clifford, M. J. Future strategies for tackling mercury pollution in the artisanal gold mining sector: Making the Minamata Convention work. Futures 2014, 62 (Part A), 106–112. 35. Hilson, G. Abatement of mercury pollution in the small-scale gold mining industry: Restructuring the policy and research agendas. Sci. Total Environ. 2006, 362, 1–14. 36. Hilson, G.; Hilson, C. J.; Pardie, S. Improving awareness of mercury pollution in small-scale gold mining communities: Challenges and ways forward in rural Ghana. Environ. Res. 2007, 103, 275–287. 37. Drace, K.; Kiefer, A. M.; Veiga, M. M.; Williams, M. K.; Ascari, B.; Knapper, K. A.; Logan, K. M.; Breslin, V. M.; Skidmore, A.; Bolt, D. A.; Geist, G.; Reddy, L.; Cizdziel, J. V. Mercury-free, small-scale artisanal gold mining in Mozambique: utilization of magnets to isolate gold at clean tech mine. J. Clean. Prod. 2012, 32, 88–95. 38. Velásquez-López, P. C.; Veiga, M. M.; Hall, K. Mercury balance in amalgamation in artisanal and small-scale gold mining: identifying strategies for reducing environmental pollution in Portovelo-Zaruma, Ecuador. J. Clean. Prod. 2010, 18, 226–232. 39. González-Carrasco, V.; Velasquez-Lopez, P. C.; Olivero-Verbel, J.; PájaroCastro, N. Air mercury contamination in the gold mining town of Portovelo, Ecuador. Bull. Environ. Contam. Toxicol. 2011, 87, 250–253. 74 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

40. Guimaraes, J. R. D.; Betancourt, O.; Miranda, M. R.; Barriga, R.; Cueva, E.; Betancourt, S. Long-range effect of cyanide on mercury methylation in a gold mining area in southern Ecuador. Sci. Total Environ. 2011, 409, 5026–5033. 41. Vieira, R. Mercury-free gold mining technologies: possibilities for adoption in the Guianas. J. Clean. Prod. 2006, 14, 448–454. 42. Veiga, M. Artisanal Gold Mining Activities in Guyana; United Nations Industrial Development Organization (UNIDO): Vienna, Austria, 1998. 43. Clifford, M. J. Pork knocking in the land of many waters: Artisanal and smallscale mining (ASM) in Guyana. Resour. Policy 2011, 36, 354–362. 44. Marsden, J. The chemistry of gold extraction, 2nd ed.; Society for Mining, Metallurgy, and Exploration: Littleton, Colo, 2006. 45. Veiga, M. M.; Baker, R. F.; Fried, M. B.; Withers, D.; Protocols for Environmental and Health Assessment of Mercury Released by Artisanal and Small-scale Gold Miners; United Nations Publications: Herndon, VA, 2004. 46. Drake, P. L.; Rojas, M.; Reh, C. M.; Mueller, C. A.; Jenkins, F. M. Occupational exposure to airborne mercury during gold mining operations near El Callao, Venezuela. Int. Arch. Occup. Environ. Health 2001, 74, 206–212. 47. ATSDR. Toxicological Profile: Mercury; http://www.atsdr.cdc.gov/ toxprofiles/tp.asp?toxid=24 (accessed Jun 27, 2015). 48. OSHA. Chemical Sampling Information. Mercury (Vapor) (as Hg); https:/ /www.osha.gov/dts/chemicalsampling/data/CH_250510.html (accessed Jun 27, 2015). 49. OSHA. Annotated PELs; https://www.osha.gov/dsg/annotated-pels/tablez2.html (accessed Jun 27, 2015). 50. CDC. Immediately Dangerous to Life or Health Concentrations (IDLH): Mercury compounds [except (organo) alkyls] (as Hg) - NIOSH Publications and Products;http://www.cdc.gov/niosh/idlh/7439976.html (accessed Jun 26, 2015). 51. Milne, J.; Christophers, A.; Silva, P. D. Acute mercurial pneumonitis. Br. J. Ind. Med. 1970, 27, 334–338. 52. Lilis, R.; Miller, A.; Lerman, Y. Acute mercury poisoning with severe chronic pulmonary manifestations. Chest 1985, 88, 306–309. 53. Hinton, J. J.; Veiga, M. M.; Beinhoff, C. Women and Artisanal Mining: Gender Roles and the Road Ahead. In The Socio-Economic Impacts of Artisanal and Small-Scale Mining in Developing Countries; Taylor and Francis e-library, 2005; pp 149–188. 54. García, O.; Veiga, M. M.; Cordy, P.; Suescún, O. E.; Molina, J. M.; Roeser, M. Artisanal gold mining in Antioquia, Colombia: a successful case of mercury reduction. J. Clean. Prod. 2015, 90, 244–252. 55. Veiga, M. M.; Meech, J. A.; Hypolito, R. Educational Measures to Address Hg Pollution from Gold Mining Activities in the Amazon. Ambio 1995, 24, 216–220. 56. Spiegel, S. J. Socioeconomic dimensions of mercury pollution abatement: Engaging artisanal mining communities in Sub-Saharan Africa. Ecol. Econ. 2009, 68, 3072–3083. 75 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

57. Spiegel, S. J.; Savornin, O.; Shoko, D.; Veiga, M. M. Mercury reduction in Munhena, Mozambique: homemade solutions and the social context for change. Int. J. Occup. Environ. Health 2006, 12, 215–221. 58. Shandro, J. A.; Veiga, M. M.; Chouinard, R. Reducing mercury pollution from artisanal gold mining in Munhena, Mozambique. J. Clean. Prod. 2009, 17, 525–532. 59. Veiga, M. Pilot Project for the Reduction of Mercury Contamination Resulting From Artisanal Gold Mining Fields in the Manica District of Mozambique; Report to the United Nations Industrial Development Organization (UNIDO) an the Blacksmith Institute, 2005; p 43. 60. Lovitz, S. B. Scales of Responsible Gold Mining: Overcoming Barriers To Cleaner Artisanal Mining In Southern Ecuador. Master of Science Specializing in Natural Resource Planning; The University of Vermont: Burlington, Vermont, 2006. 61. Baeuml, J.; Bose-O’Reilly, S.; Lettmeier, B.; Maydl, A.; Messerer, K.; Roider, G.; Drasch, G.; Siebert, U. Applicability of two mobile analysers for mercury in urine in small-scale gold mining areas. Int. J. Hyg. Environ. Health 2011, 215, 64–67. 62. Pogarev, S. E.; Ryzhov, V.; Mashyanov, N.; Sholupov, S.; Zharskaya, V. Direct measurement of the mercury content of exhaled air: a new approach for determination of the mercury dose received. Anal. Bioanal. Chem. 2002, 374, 1039–1044. 63. Dondeyne, S.; Ndunguru, E.; Rafael, P.; Bannerman, J. Artisanal mining in central Mozambique: Policy and environmental issues of concern. Resour. Policy 2009, 34 (1-2), 45–50. 64. Dondeyne, S.; Ndunguru, E. Artisanal gold mining and rural development policies in Mozambique: Perspectives for the future. Futures 2014, 62 (Part A), 120–127. 65. Veiga, M. M.; Nunes, D.; Klein, B.; Shandro, J. A.; Velasquez, P. C.; Sousa, R. N. Mill leaching: a viable substitute for mercury amalgamation in the artisanal gold mining sector? J. Clean. Prod. 2009, 17, 1373–1381. 66. Steckling, N.; Bose-O’Reilly, S.; Shoko, D.; Muschack, S.; Schierl, R. Testing Local Conditions for the Introduction of a Mercury-free Gold Extraction Method using Borax in Zimbabwe. J. Health Pollut. 2014, 4, 54–61. 67. Kiefer, A. M.; Drace, K.; Gottlieb, S.; Coursey, S.; Veiga, M. M.; da Cruz Marrumbe, P. N.; Jose Chapo, M. A. Evaluation of mercury content in amalgams from Munhena mine, Mozambique. J. Clean. Prod. 2014, 84, 783–785. 68. Sandoval, F. Small-scale Mining in Ecuador; Mining, Minerals and Sustainable Development 75; International Institute for Environment and Development: London, UK, 2001. 69. Grylls, C. There’s gold in that there hill. Focus. September 1998; pp 114–118. 70. Ramırez Requelme, M. E.; Ramos, J. F. F.; Angélica, R. S.; Brabo, E. S. Assessment of Hg-contamination in soils and stream sediments in the 76 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

71. 72.

73.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch003

74.

75.

76.

77.

mineral district of Nambija, Ecuadorian Amazon (example of an impacted area affected by artisanal gold mining). Appl. Geochem. 2003, 18, 371–381. Dalgety, W. T. Placer mining and the Guyana environment; Paper Craft: Georgetown, Guyana, 2010. Hennessy, L. Where There Is No Company: Indigenous Peoples, Sustainability, and the Challenges of Mid-Stream Mining Reforms in Guyana’s Small-Scale Gold Sector. New Polit. Econ. 2015, 20, 126–153. Miller, J. R.; Lechler, P. J.; Bridge, G. Mercury Contamination of Alluvial Sediments within the Essequibo and Mazaruni River Basins, Guyana. Water, Air, Soil Pollut. 2004, 148, 139–166. Hammond, D. S.; Gond, V.; Thoisy, B. de; Forget, P.-M.; DeDijn, B. P. E. Causes and Consequences of a Tropical Forest Gold Rush in the Guiana Shield, South America. Ambio 2007, 36, 661–670. Howard, J.; Trotz, M. A.; Thomas, K.; Omisca, E.; Chiu, H. T.; Halfhide, T.; Akiwumi, F.; Michael, R.; Stuart, A. L. Total mercury loadings in sediment from gold mining and conservation areas in Guyana. Environ. Monit. Assess. 2011, 179, 555–573. Adler Miserendino, R.; Bergquist, B. A.; Adler, S. E.; Guimarães, J. R. D.; Lees, P. S. J.; Niquen, W.; Velasquez-López, P. C.; Veiga, M. M. Challenges to measuring, monitoring, and addressing the cumulative impacts of artisanal and small-scale gold mining in Ecuador. Resour. Policy 2013, 38, 713–722. Metcalf, S. M.; Veiga, M. M. Using street theatre to increase awareness of and reduce mercury pollution in the artisanal gold mining sector: a case from Zimbabwe. J. Clean. Prod. 2012, 37, 179–184.

77 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Chapter 4

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

Tracing Dust Sources Using Stable Lead and Strontium Isotopes in Central Asia Nitika Dewan,1 Brian J. Majestic,*,1 Michael E. Ketterer,2 Justin P. Miller-Schulze,3 Martin M. Shafer,4,5 James J. Schauer,4,5 Paul A. Solomon,6 Maria Artamonova,7 Boris B. Chen,8 Sanjar A. Imashev,8 and Gregory R. Carmichael9 1Department

of Chemistry and Biochemistry, University of Denver, Denver, Colorado 80208, U.S.A. 2Department of Chemistry, Metropolitan State University of Denver, 1201 5th Street, Denver, Colorado 80204, U.S.A. 3Department of Chemistry, 6000 J Street, California State University, Sacramento, California 95819, U.S.A. 4Wisconsin State Laboratory of Hygiene, 2601 Agriculture Drive, Madison, Wisconsin 53718, U.S.A. 5Environmental Chemistry and Technology Program, 660 North Park Street, University of Wisconsin, Madison, Wisconsin 53706, U.S.A. 6U.S. EPA, Office of Research and Development, Las Vegas, Nevada 89193, U.S.A. 7Institute of Atmospheric Physics, 109017 Moscow, Russia 8Kyrgyz-Russian Slavic University, 44 Kievskaya Street, Bishkek 720000, Kyrgyzstan 9Department of Chemical and Biochemical Engineering, The University of Iowa, Iowa City, Iowa 52242, U.S.A. *E-mail: [email protected].

From 1960 to 2014, the Aral Sea’s surface area has receded about 90% in size from 68,000 km2 to 8,444 km2. Consequently, newly exposed sediments are resuspended by the wind and are now a source of atmospheric particulate matter in Central Asia, which may have an impact on human health and climate. In this study, strontium (Sr) and lead (Pb) stable isotopic ratios, along with other elemental compositions, are used to determine if the Aral Sea sediments are an important source of © 2015 American Chemical Society In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

air pollution to Central Asia. Ambient particulate matter with aerodynamic diameter < 10 µm (PM10) samples were collected every other day and included dust and non-dust events at the Bishkek and LIDAR (in Teploklyuchenka) in Kyrgyzstan. Soil samples also were collected in the vicinity of the air sampling sites, resuspended, and sized as PM10 for chemical analysis. The average 87Sr/86Sr ratio for the Aral Sea sediments was 0.70992 (range, 0.70951 - 0.71064), which is less radiogenic than the surface soils in Kyrgyzstan showing an average ratio of 0.71579 (range, 0.71448 - 0.71739). In contrast, the airborne PM10 collected in Kyrgyzstan had an average 87Sr/86Sr ratio of 0.71177 (range, 0.70946 - 0.71335), which is between the two ratios, indicating a possible mixture of sources. However, no differences in Sr ratios were observed between dust and non-dust events, which implies that the impact of Aral Sea sediments on the sampling sites is minimal. The element enrichment factors and stable Pb isotope ratios are employed to further understand the source of PM10. Airborne PM10 are characterized by enrichments in elements like As, Cd, Cu, Mo, Pb, and Zn at both sampling sites. The mean (K/Pb) ratio for aerosols is ~ 45 and for soils is ~800, which suggest that the aerosols contain a significant fraction of anthropogenic source of airborne Pb in Kyrgyzstan. 208Pb/206Pb ratios are higher for aerosols compared to the Kyrgyzstan soils and Aral Sea sediments also suggesting that Pb is most likely present due to anthropogenic sources.

Introduction Atmospheric particulate matter (PM) is a mixture of solid particles and liquid droplets suspended in the air. This includes organic carbon, elemental carbon, anions, cations, bulk metals, and trace elements. These particles exist in different shapes, concentrations, and are of varying chemical composition depending upon origin, source of emissions, atmospheric processing, and season. The United States Environmental Protection Agency regulates PM in two categories based on sizes e.g., fine particles (PM2.5) are less than 2.5 µm in diameter and coarse particles (PM10) which are less than 10 µm in diameter. PM in the atmosphere is a serious issue because of its impact on climate, human health, visibility, and atmospheric reactivity (1, 2). PM can travel up to thousands of km and their effects can be observed downwind, not just at the site of the emission (3). For example, dust from the Gobi Desert in Western China travels more than 10,000 km, affecting the air quality on the western coast of the United States (4). As particles can be transported, it is essential to determine the source of their emission. Recently, studies have shown that stable isotope ratios (e.g., Sr and Pb) are potent tools for source origin (5, 6) and tracing dust sources (7, 8). The power 80 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

of Pb and Sr isotopic system lies in the fact that biological processes and phase changes do not fractionate these isotopes; therefore, the measured ratio represents the exact source or mixture of sources (9). Both rocks and minerals possess unique 87Sr/86Sr ratios depending upon the geological origin (10). 87Sr is a radiogenic and non-radioactive isotope produced from the radioactive decay of 87Rb; therefore, increased 87Sr/86Sr in older rocks and sediments is observed (10). Stable Pb isotope ratios allow efficient tracing of the sources of Pb pollution in the atmosphere (11). Pb has three naturally occurring stable radiogenic isotopes; 206Pb, 207Pb, and 208Pb, produced from the radioactive decay chain of 238U, 235U, and 232Th respectively. The only stable non-radiogenic isotope is 204Pb. 207Pb/206Pb, 208Pb/204Pb, 206Pb/ 204Pb, and 207Pb/204Pb ratios are useful in differentiating anthropogenic sources from natural sources (11, 12). These ratios can vary considerably depending on their origin, rainfall, wind direction, and the presence or absence of anthropogenic sources (13, 14). The Aral Sea, once one of the largest lakes of the world, is located in Central Asia between Kazakhstan and Uzbekistan. The Aral Sea’s surface area has receded 90% in size due to removal of water for irrigation, thereby transforming the region into a salt desert (15). Several ecological problems like soil salinization, groundwater table reduction (16), and increase in intensity and frequency of dust storms have resulted due to the Aral Sea’s desiccation (17). The Aral Sea basin also serves as a trap for wind-blown dust originating from deserts on all sides: Circum-Aral Karakum to the north, Kyzylkum to the east, Karakum to the south, and Ustyurt to the west, thus dust storms originating in the Aral Sea likely contain a mixture of these deserts, in addition to the native sediments (18). These relatively recently exposed sediments, emitted into the air, may be a previously unidentified source of PM10 in Central Asia, which could have an impact on human health and climate (19). In this study, we aim to characterize sources of PM10 from two sites in Kyrgyzstan to advance our understanding of regional and long-range transport of aerosol PM. Herein, we expand upon data and conclusions previously published in Atmospheric Environment (Atmos. Environ.) (20). We focus on the Aral Sea region, which is now the third largest source of mineral dust in Asia (21). The primary goal of this study is to determine if the resuspended Aral Sea sediments are an important source of air pollution in Kyrgyzstan. This is achieved by examining differences between the metal content of PM10 from dust event and non-dust event periods at both sampling sites and whether the Sr and Pb isotopic composition of the Aral Sea sediments and local soils collected in Kyrgyzstan are distinguishable from each other and those of previously published Western and Central China data.

Experimental Sample Collection Sites Kyrgyzstan is a country located in Central Asia, which is bordered by Kazakhstan to the north, Tajikistan to the south, China to the east, and Uzbekistan to the west. Bishkek is the capital and the largest city of Kyrgyzstan with a 81 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

population of ~854,000 (20). Kyrgyzstan is a mountainous country dominated by the Tien-Shan mountain range encompassing the entire nation. The samples were collected at two urban sites in Kyrgyzstan: Bishkek and LIDAR. The Bishkek site (42° 40′ 47.80″ N, 74° 31′ 44.30" E, altitude 1,250 m) is located ~23 km south of the Bishkek city center. The LIDAR site (42° 27′ 49.30″ N, 78° 3123 km south of the Bishkek city center. The LIDAR site (42° 27′ 49.30″ N, 78° 49.30″ E, altitude 1920 m) is located in Teploklyuchenka, which is about 380 km east of the Bishkek site. Both sampling sites are located in the mountains and the distance directly between them is approximately 315 km. The Bishkek site is ~ 1,200 km and the LIDAR site is ~1,500 km east-southeast (ESE) from the Aral Sea. A map of the region and sampling sites in shown in Figure 1 (20).

Figure 1. Map of Central Asia and geographical location of the sampling sites (shown as stars) and surrounding deserts (1-Kyzylkum, 2-Aral Karakum, 3-Karakum, 4-Ustyurt and Mangyshlak, 5-Betpak Dala, 6-Saryesik-Atyrau Desert, 7-Taukum, 8-Qaratal and Lepsy, 9-Moinkum, 10-Aralkum) relative to Aral Sea location. Reproduced with permission from reference (20). Copyright (2015) Elsevier.

Sample Collection Sediments collected at the Aral Sea and Kyrgyzstan soils collected at the Bishkek and LIDAR sites were resuspended in the laboratory and collected on filters as PM10 (particles < 10 µm in diameter) for subsequent analysis. In addition, atmospheric PM10 samples were collected every other day at the Kyrgyzstan sites from mid-July 2008 to mid-July 2009 during dust and non-dust events. The 82 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

objective of this study was to focus on the long-range and trans-boundary transport of pollution from outside and within Central Asia, respectively; therefore, to minimize the influence of local PM10 sources, these sites were selected based on their distance from more populated areas. The PM10 samples were collected on pre-cleaned 47 mm Teflon filters (Teflo, Pall-Gelman) using URG 3000 ABC samplers (URG Corporation, U.S.A.) at a flow rate of 8 L min–1 for 24 h at each sampling site. Filters and filter portions were composited into biweekly or monthly samples depending on the sample loading. Additional details regarding sampling sites and sample collection can be found elsewhere (19). Chemical Analysis Sample preparation was performed under positive pressure HEPA filtered air. Using a high-precision microbalance (MX5, Mettler-Toledo, U.S.A.), the mass of the collected PM10 was determined gravimetrically by weighing the filters that were equilibrated pre- and post-sampling at constant humidity (35 ± 3%) and temperature (21 ± 2 °C) for 24 h. The mass measurement had an uncertainty of < 7% or ± 4 µg. A Po-ionization source was used to remove the static charges on the filters before weighing. The airborne PM10 and resuspended PM10 Kyrgyzstan soils and Aral Sea sediments were digested in a 36-position Microwave Rotor (Milestone Ethos). The digestion matrix consisted of 1 mL nitric (16 M), 0.25 mL hydrochloric (12 M), and 0.1 mL hydrofluoric acid (28 M). The microwave digested samples were diluted to 15 mL with Millipore water (>18 MΩcm, MQ) and elemental concentrations were determined using high-resolution magnetic sector inductively coupled plasma mass spectrometry (HR-ICP-MS, Element 2, Thermo-Fisher). Laboratory blanks and field blanks were rigorously applied to account for potential contamination of filters and reagents. Uncertainty for each element was determined from error propagation analysis. Analysis of Sr Stable Isotope Ratios Prior to Sr isotope analysis, Rb was removed since 87Rb interferes with the Sr isotope measurement. In this multistep process, the samples were first evaporated in Teflon vials and diluted to 1 mL with 2 M optima grade nitric acid (Fisher, U.S.A.). In the digests, Rb content was about 4 - 180 ng and Sr content was about 18 - 680 ng. Several airborne PM10 samples were composited to ensure concentrations above the method detection limit of 18 ng for Sr. Sr was separated from Rb using Sr Spec resin (Eichrom, U.S.A.) (22). The Sr Spec resin slurry was prepared in 0.05 M HNO3 and loaded into modified 5 mL glass wool fitted plastic Pasteur pipettes. The Sr Spec column was cleaned by passing 3 × 5 mL Millipore water, 2 mL 0.05 M HNO3, 5 mL of Millipore water through the column. The column was then conditioned by passing 4 × 0.5 mL 2 M HNO3 through it. Following column conditioning, the sample in 2 M HNO3, was loaded in 2 × 0.5 mL aliquots. The sample loaded column was washed with 1 mL of 2 M HNO3, 8 × 0.5 mL of 7 M HNO3, and 1 mL of 3 M HNO3, while Sr was retained. Elution of Sr was accomplished by passing 6 × 0.5 mL of 0.05 M HNO3 through the column. 83 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

Sr concentrations were measured using a quadrupole inductively coupled plasma mass spectrometry (ICP-MS, Agilent 7700). The Sr recoveries for the samples were between 90-110%. 87Sr/86Sr ratios of the Aral Sea sediments were measured by a Nu Plasma II multi-collector inductively coupled plasma mass spectrometer (MC-ICP-MS) and the PM10 and Kyrgyzstan soils by a Thermo-Finnigan Neptune MC-ICP-MS. Stable Pb isotopic ratios were measured in extracts with no further purification by using high-resolution magnetic sector inductively coupled plasma mass spectrometer (Thermo-Finnigan Neptune Plus). The Sr and Pb chemical separation of the samples was validated by analyzing certified standard references, NBS 987 and NIST 981, respectively. The NBS 987 yielded 87Sr/86Sr = 0.71030 ± 0.00001 (n = 17) by Neptune and 87Sr/86Sr =0.71023 ± 0.00003 (n = 17) by Nu Plasma. The certified value of 87Sr/86Sr for NBS 987 is 0.71024 ± 0.00007. The NIST 981 yielded 208Pb/206Pb = 2.1656 ± 0.0014 (n = 16) and 207Pb/206Pb = 0.9144 ± 0.0001 (n = 16). The certified value of 208Pb/206Pb for NIST 981 is 2.1681 ± 0.0008 and 207Pb/206Pb is 0.9146 ± 0.0003.

Results and Discussion Site-Dependent Elemental Concentrations Trace and bulk elements were quantified in the resuspended Kyrgyzstan soils, resuspended Aral Sea sediments, and airborne PM10 collected at the two sampling sites during dust events and non-dust events (Tables 1 and 2). The major crustal element levels (µg g–1) of the LIDAR soils were Al (~67000 ± 6000), Fe (~50000 ± 3500), and Ca (~29000 ± 2500); and of Bishkek soils were Al (~63000 ± 6000), Fe (~40000 ± 3000), and Ca (~22000 ± 2000). This indicates that the elemental composition of the Kyrgyzstan soils at both sites were similar. In contrast, in the resuspended Aral Sea sediments, Ca (~50000 ± 3500) was the dominant element followed by Fe (~20000 ± 1500), and Al (~23000 ± 1500). Approximately, two or three fold differences also existed in the concentrations of Na, Cr, S, Ti, Cu, Ni, Zn, Rb, Sr, Mo, and Sb in Aral Sea sediments compared to the Kyrgyzstan soils (Table 1). The elements measured in PM10 showed significantly higher concentrations during dust events compared to non-dust events at the sampling sites (Table 2). For instance, concentrations (µg g–1) of the crustal related elements, such as Al, Ca, and Fe were ~11300, ~18000, and ~9800 respectively, during dust events and ~9900, ~13000, and ~9000 during non-dust events at the Bishkek site. Similarly, at the LIDAR site, Al, Ca, and Fe concentrations (µg g–1) were ~20000, ~26000, and ~19000, respectively during dust events and ~11000, ~15000, and ~12000 during non-dust events. At both sites, the dominant element in PM10 was Ca followed by Fe, K, and Al. Overall, the concentrations of most of the elements were higher at the LIDAR site relative to the Bishkek site. This may be because the LIDAR site is much closer to the Taklamakan Desert and thus more impacted by the dust transport from this desert (23). 84 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

Table 1. Average Elemental Concentrations (µg g–1) of Resuspended Kyrgyzstan Soils and Resuspended Aral Sea Sediments. Reproduced with permission from reference (20). Copyright (2015) Elsevier. Element

Bishkek soils (µg g–1)

LIDAR soils (µg g–1)

Al

63000 ± 6000

67000 ± 6000

23000 ± 1500

K

19000 ± 6000

21000 ± 7000

16000 ± 3000

Ca

22000 ± 2000

29000 ± 2500

50000 ± 3500

Mg

10000 ± 1000

15000 ± 1100

6000 ± 500

Na

12000 ± 1100

16000 ± 1500

9500 ± 750

Cr

90 ± 6

100 ± 6

250 ± 40

P

600 ± 50

1100 ± 60

500 ± 30

Mn

800 ± 60

960 ± 60

650 ± 30

Fe

40000 ± 3000

50000 ± 3500

20000 ± 1500

Ni

42 ± 3

43 ± 4

60 ± 6

V

120 ± 7

130 ± 8

100 ± 5

Zn

56 ± 6

60 ± 7

73 ± 9

Co

16 ± 1

17 ± 1

11 ± 1

Cu

32 ± 2

34 ± 2

21 ± 1

Ti

4600 ± 360

5300 ± 400

1800 ± 120

As

11 ± 3

9±2

15 ± 3

Rb

60 ± 5

81 ± 6

46 ± 5

Sr

160 ± 15

250 ± 20

630 ± 50

Mo

1.0 ± 0.1

2.0 ± 0.2

8±1

S

370 ± 30

480 ± 40

11000 ± 1000

Sb

1.5 ± 0.1

1.2 ± 0.1

25 ± 2

Ba

500 ± 40

530 ± 40

400 ± 20

Pb

24 ± 2

25 ± 2

18 ± 2

La

20 ± 1

30 ± 2

15 ± 1

Th

8±1

9±2

5.0 ± 0.7

U

4.0 ± 0.3

9±1

3.0 ± 0.3

Aral Sea sediments (µg g–1)

Continued on next page.

85 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

Table 1. (Continued). Average Elemental Concentrations (µg g–1) of Resuspended Kyrgyzstan Soils and Resuspended Aral Sea Sediments. Element

Bishkek soils (µg g–1)

LIDAR soils (µg g–1)

Sn

4.5 ± 0.3

3.7 ± 0.3

1.2 ± 0.5

K/Pb

775

863

842

Aral Sea sediments (µg g–1)

Table 2. Average Elemental Concentrations (µg g–1) of Airborne PM10 Samples Collected at the Bishkek and LIDAR Sampling Sites Element

Bishkek dust events PM10 (µg g–1)

Bishkek non-dust events PM10 (µg g–1)

LIDAR dust events PM10 (µg g–1)

LIDAR non-dust events PM10 (µg g–1)

Al

11300 ± 1050

9900 ± 100

20000 ± 1800

11000 ± 1000

K

9000 ± 120

8000 ± 1000

16000 ± 3000

14000 ± 2500

Ca

18000 ± 1600

13000 ± 1200

26000 ± 2000

15000 ± 1200

Mg

3500 ± 250

3000 ± 300

7000 ± 500

4000 ± 300

Na

5000 ± 450

3500 ± 350

7000 ± 300

4500 ± 400

Cr

30 ± 5

37 ± 5

44 ± 2

66 ± 7

P

25 ± 2

10 ± 1

15 ± 1

10 ± 1

Mn

280 ± 18

260 ± 15

480 ± 30

350 ± 20

Fe

9800 ± 100

9000 ± 600

19000 ± 1200

12000 ± 1000

Ni

25 ± 8

25 ± 7

31 ± 5

30 ± 5

V

57 ± 3

65± 4

73 ± 5

65 ± 5

Zn

230 ± 30

300 ± 40

150 ± 30

330 ± 85

Co

50 ± 11

64 ± 15

350 ± 30

150 ± 15

Cu

45 ± 8

103 ± 12

320 ± 30

100 ± 10

Ti

1200 ± 100

1100 ± 100

2000 ± 150

1500 ± 120

As

24 ± 5

30 ± 5

21 ± 4

25 ± 5

Rb

25 ± 3

25 ± 2

55 ± 6

35 ± 7

Sr

180 ± 18

110 ± 10

200 ± 18

130 ± 12

Mo

2.5 ± 0.2

3.3 ± 0.4

2.6 ± 0.3

4±1

S

15000 ± 1500

17000 ± 1600

17000 ± 1500

10000 ± 1000

Continued on next page.

86 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

Table 2. (Continued). Average Elemental Concentrations (µg g–1) of Airborne PM10 Samples Collected at the Bishkek and LIDAR Sampling Sites Element

Bishkek dust events PM10 (µg g–1)

Bishkek non-dust events PM10 (µg g–1)

LIDAR dust events PM10 (µg g–1)

LIDAR non-dust events PM10 (µg g–1)

Sb

15 ± 1

15 ± 1

8±1

13 ± 2

Ba

150 ± 13

150 ± 10

270 ± 20

200 ± 15

Pb

180 ± 30

230 ± 40

160 ± 15

100 ± 12

La

5±1

5.5 ± 0.4

12 ± 3

7±1

Th

1.8 ± 0.1

2.0 ± 0.2

4±1

3±1

U

2.1 ± 0.2

1.9 ± 0.1

3.0 ± 0.5

2.3 ± 0.1

Sn

9±3

20 ± 3

5±3

10 ± 3

K/Pb

57

46

112

76

Enrichment Factor The enrichment factor (EF) of the elements was calculated by first normalizing the elemental concentrations in the sample with aluminum (Al) (24), and then dividing by the Upper Continental Crust (UCC) ratio (25). The EF is used to identify anthropogenic components of aerosols in the atmosphere. EF is close to unity for the elements related to the reference, Al (marker for crustal emissions). A high EF (>> 10) suggests an important anthropogenic source is associated with that pollutant (24). EF can be calculated using the following formula

The dashed line (EF = 10) on the plots shown in Figure 2(a) and 2(b) represents the level above which the element is considered to be anthropogenically sourced (24). There is no difference in EF for crustal elements like Ca, Fe, K, Mg, Na, and Ti during dust and non-dust events at both the sampling sites. Also, these elements were associated with natural sources at the Bishkek and LIDAR sites. The toxic elements like As, Cd, Cr, Cu, Mo, Sb, Sn, Pb, and Zn were highly enriched at both sampling sites. In Figure 2(a), significant differences were observed between the dust and non-dust events, with EF higher during the non-dust events at the LIDAR sites. The implication here is that anthropogenic sources are dominant during the non-dust events. At the Bishkek site, however, there was very little difference in EF between dust and non-dust sampling periods, aside from Cu and Sn as shown in Figure 2(b). 87 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

Figure 2. Enrichment Factor (EF) for PM10 collected at 2(a) LIDAR site during dust and non-dust events and 2(b) Bishkek site during dust and non-dust events.

88 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

Figure 3. Correlation between elemental Pb (µg g–1) and elemental Th (µg g–1) for 3(a) Aral Sea sediments and Kyrgyzstan soils and 3(b) PM10 during dust and non-dust events at the two sampling sites. Reproduced with permission from reference (20). Copyright (2015) Elsevier.

89 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

208Pb

is the final step in the radioactive decay chain of 232Th (26). Anthropogenic Pb sources from Aral Sea sediments, local soils, and airborne PM10 can be differentiated from natural Pb sources by plotting concentrations (µg g–1) of elemental Pb against elemental Th (20, 27). In Figure 3(a), thorium (Th) correlates linearly with Pb for Aral Sea sediments and Kyrgyzstan soils, indicating that Pb in these samples is primarily from natural sources (27). This is also consistent with low Upper Continental Crust Enrichment Factor for Pb for both soils (EFPb = 1.54) and Aral Sea sediments (EFPb = 3.09), since EF is less than 10. The K/Pb ratio for the local soils and Aral Sea sediments is in between 775 - 863, as shown in Table 1, which also implies that the Pb in these samples is most likely of natural origin (28, 29). The lack of relationship [Figure 3(b)] between the elemental Th and elemental Pb in PM10 indicates that airborne Pb is dominated by anthropogenic sources and not dust sources, such as from Aral Sea or local soils. This observation was in agreement with high UCC EFs for PM10 (EFPb = 37) at Bishkek site and (EFPb = 74) at LIDAR site. Also, the range of K/Pb ratios is in between 46-112, as shown in Table 2, indicating an anthropogenic component of Pb in PM10 at both sampling sites.

Impact of the Aral Sea The 87Sr/86Sr ratios for Aral Sea sediments, Kyrgyzstan soils, and airborne PM10 collected at Bishkek and LIDAR sites during dust and non-dust events as presented in the Atmos. Environ. Paper (20). The 87Sr/86Sr ratios for the resuspended Aral Sea sediments are between the range 0.70951 - 0.71064, whereas the local soils ratio are between the range 0.71448 - 0.71739. The PM10 ratios are in the range of 0.70946 - 0.71335. The 87Sr/86Sr ratios of airborne PM10 mainly fall between these potential sources. To determine the dust sources in PM10, 206Pb/204Pb is plotted against 87Sr/86Sr in Figure 4 (30). In the 87Sr/86Sr domain, a mixture consisting of only two sources (e.g., Aral Sea sediments and Kyrgyzstan soils) will lie on straight line. If not, additional sources are likely present. The data can also be plotted using Sr isotopic ratio against 1/Sr concentration (µg g–1) (20). The Aral Sea sediments are encircled and Kyrgyzstan soils are inside the square. The average 87Sr/86Sr ratio for the Aral Sea sediments and surface soils in Kyrgyzstan were 0.70992 and 0.71579, respectively, allowing differentiation between these two source types. The average Sr ratio for soils at Bishkek and LIDAR were similar (p > 0.05) and statistically different than the soils collected near the sites (p < 0.05) (20). The average 87Sr/86Sr ratio for the atmospheric PM10 at Bishkek site was 0.71047 (range, 0.70946 - 0.71156) and 0.71104 (range, 0.71027 - 0.71218) during dust and non-dust events, respectively. Similarly, the average 87Sr/86Sr ratio at LIDAR site was 0.71240 (range, 0.71088 - 0.71335) and 0.71179 (range, 0.71100 - 0.71285) during dust events and non-dust events, respectively. Ratios for the dust events and non-dust events are not statistically different at either site (t-test, p > 0.05). Although the dust and non-dust ratios are not significantly different, Sr ratios observed in PM10 (particularly at Bishkek), are more similar to 90 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

the Aral Sea sediments or perhaps other deserts in the region relative to the local soils suggesting that local soils may not be an important source of PM10 at these locations. Based on the back trajectory analysis using HYSPLIT (Hybrid Single Particle Lagrangian Integrated Trajectory), the Aral Sea may have had an impact on the dust events at the Bishkek and LIDAR sites. The impacted samples are indicated by arrows in Figure 4. The model also showed the impact of Algeria, Gobi Desert, Iran, Libya, and Mediterranean Sea on the dust events at the two sites. There is no statistical difference in the 87Sr/86Sr ratios between dust events that passed over Aral Sea and the ones that took different trajectory suggesting that the Aral Sea sediments, at best, have only a minor influence in Kyrgyzstan and dust from other regions as well as regional anthropogenic sources that may be impacting PM10 concentrations in Kyrgyzstan.

Figure 4. 206Pb/204Pb vs. 87Sr/86Sr for Aral Sea sediments (in circle), local soils (in square), and PM10. The arrows suggest the impact of Aral Sea on dust events at Bishkek and LIDAR sites based on HYSPLIT modelling.

Impact of Other Source Regions Since the results above suggest that the Aral Sea appears to have only a minimal effect (at best) on the air quality in Kyrgyzstan, we compared our isotopic results with those from other potential source regions. To evaluate this, 87Sr/86Sr is plotted against 87Rb/86Sr ratios for Aral Sea sediments, Kyrgyzstan soils, and airborne PM10, with those of sediments from Western China (31) and the Tien-Shan mountain range bordering Kyrgyzstan (Figure 5).20 The 87Sr/86Sr 91 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

ratios for PM10 are almost identical to the ratios observed for sediments from North Tien-Shan region, which is a large mountain range located in Central Asia (32). These results suggest that transport from neighboring regions to Kyrgyzstan is important source of dust. The Sr ratios for PM10 were also similar to some of the soils measured in Tarim Basin, a major source of dust in Western China (32). However, previous modelling studies in the region suggest that trajectories may have also originated due to long-range transport from Africa, or Middle East regions (23, 33, 34). Sr isotopic ratio from these regions would be needed, however, to evaluate the impact of long-range transport as a significant contributor of dust to Kyrgyzstan.

Figure 5. Comparison of isotopic composition of the Aral Sea sediments, local soils, and PM10 with those of soils from Western and Central China.

Pb Isotope Ratios Stable Pb isotope ratios (208Pb/204Pb, 206Pb/204Pb, 206Pb/207Pb, and for resuspended Aral Sea sediments and local soils, and airborne PM10 are presented elsewhere (20). Ratios of 208Pb/206Pb versus 206Pb/207Pb are presented in Figure 6 for the Aral Sea sediments (in circle), Kyrgyzstan soils (in dotted circle) and PM10 (in rectangle). The average 208Pb/206Pb ratios of the Aral Sea sediments and Kyrgyzstan soils are 2.095 (range, 2.090 - 2.098) and 2.069 (range, 1.994 - 2.071), respectively. These ratios also suggest that both sediments and soils have a natural Pb source (35), which is also in agreement with the low UCC EF (range, 1.54 – 3.09) and high K/Pb ratio (range, 775 – 863). The average 208Pb/206Pb ratio for PM10 at the Bishkek site was 2.107 (range, 2.101 - 2.116) and 2.104 (range, 2.102 - 2.107) during dust events and non-dust events, respectively. Similarly, the average 208Pb/206Pb ratio at the LIDAR site was 2.099 (range, 2.091 - 2.108) and 2.108 (range, 2.102 - 2.112) during dust events 208Pb/206Pb)

92 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

and non-dust events, respectively. The average ratio for PM10 is 2.104 (range, 2.091 - 2.112), which is higher than both soils and the Aral Sea sediments. It is important to note that dust and non-dust events are not statistically different, also suggesting that airborne PM10 most likely impacted by anthropogenic Pb sources, which is consistent with high UCC EF (range, 37 - 74) and low K/Pb ratios (range, 46 – 112). Further, Figure 6 indicates that PM10 is not, as originally hypothesized, mostly a mixture of local soils and Aral Sea sediments, which is in agreement with the Sr results, all suggesting that one or more other Pb sources impact PM10 concentrations in the area. The data are plotted using another isotopic ratio in the Atmos. Environ. paper showing similar conclusions and make the conclusions more robust (20).

Figure 6. 208Pb/206Pb and 206Pb/207Pb in resuspended Kyrgyzstan soils, resuspended Aral Sea sediments, and PM10 during dust events and non-dust events at the two sampling sites.

Comparing Pb Isotopic Ratios with Other Regional Measurements Figure 6 indicates that an unknown source of Pb is impacting ambient particle concentrations in the area. Therefore, similar to Sr system, we compare the Pb isotope results to results from other regional studies so that possible sources can be identified. Figure 7 shows 208Pb/204Pb plotted against 206Pb/204Pb ratios for the Aral Sea sediments, Kyrgyzstan soils and airborne PM10, with those of total suspended particles (TSP) and soils from Western China. The PM10 Pb ratios are similar to the Pb ratios observed in TSP (particles with aerodynamic diameter < 93 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

100 µm) collected on Teflon filters over the intervals of 1-7 days from May 2001 to September 2001 using a high volume bulk aerosol sampler operated at a flow rate of 16.7 L min–1 in Kosan and Dunhuang in South Korea (36) and in Western China (31). This implies a similar Pb source(s) between the sampling sites in this study, Dunhuang and Kosan, but the source has not been identified as yet. Back trajectory analysis of pollution to Dunhuang and Kosan is from Central Asia, suggesting that the Pb in PM10, is in fact, anthropogenic and likely includes local and regional sources and sources transported outside the region (Taklamakan Desert from Western China, Europe, North Africa, parts of Russia) (20). As noted above, this is further suggested by the low K/Pb ratio and high enrichment factors observed in this study. Results in Figure 7 also indicate that the Aral Sea sediments have a similar Pb source relative to other Asian soils, including those near the sites, N. Pacific Dust and Asian Dust (31, 36). Since composition data from most other deserts in Central Asia and the surrounding regions are not available, these results also suggest that Kyrgyzstan is likely impacted by dust from these other deserts supporting above results that the Aral Sea only has a minor impact at the sampling sites (20).

Figure 7. Comparison of isotopic composition of the Aral Sea sediments, local soils, and PM10 with those of soils from Western China.

Conclusions The initial hypothesis of this study was that newly exposed Aral Sea sediments might be a new and important source of dust in Kyrgyzstan. Based on elemental concentrations, the chemical composition of LIDAR and Bishkek soils were similar. The airborne PM10 collected at two sampling sites was also similar. The Aral Sea sediment’s elemental composition was different from both soils and PM10 for majority of the elements. Also, the composition of soils were different from PM10. This further supports that local soils had no impact and the 94 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

Aral Sea had minimal impact on PM10 in Kyrgyzstan as shown in Table 1-2. Sr and Pb isotope analysis indicated that the impact of the Aral Sea is small and cannot be quantified based on the approach used in this study; although, other approaches arrived at the same conclusion (19, 23, 33, 34). Although previous studies do suggest long-range transport to be important (19, 34) our Sr isotope ratios suggest that dust from the more local Tien-Shan mountain range may also be significant. The Pb isotope ratios suggest that both Aral Sea sediments and Kyrgyzstan soils have natural source of Pb whereas airborne PM10 is dominated by unidentified combustion and/or anthropogenic sources. The similarities from these data compared to those collected in Dunhuang and Kosan hint at a similar source. We hypothesize that this source may be transport from industrial Europe, parts of Russia, or the Middle East (34). Both Sr and Pb isotope systems are consistent indicating little if any contribution from Aral Sea sediments and/or local soils to PM10 levels during dust events at Kyrgyzstan. Based on this and related work, Central Asia is impacted by yet unidentified regional and distance dust and anthropogenic sources that requires additional measurements of the comparison of desert sands in and around Central Asia as well as a network of PM10 and PM2.5 chemical speciation monitoring to understand both the sources to ambient PM and impact of transport of PM from this region.

Acknowledgments The authors thank Dr. Gwyneth Gordon and Dr. Rasmus Andreason for their help and guidance in Sr ratio measurements. The authors also gratefully acknowledge the NOAA Air Resources Laboratory (ARL) for use of the HYSPLIT transport and dispersion model (www.ready.noaa.gov). We acknowledge the Association of Public Health Laboratories (APHL) for funding through their Environmental Health Fellows program. The US Environmental Protection Agency, through its Office of Research and Development, funded this study and collaborated in the research described here under Contract EP-D-06-001 to the University of Wisconsin-Madison as a component of the International Science & Technology Center (ISTC) project # 3715 (Transcontinental Transport of Air Pollution from Central Asia to the US). The isotope portion of this study was funded through a PROF Grant at the University of Denver.

References 1. 2.

Davidson, C. I.; Phalen, R. F.; Solomon, P. A. Airborne Particulate Matter and Human Health: A review. Aerosol Sci. Technol. 2005, 39, 737–749. Solomon, P. A.; Costantini, M.; Grahame, T. J.; Gerlofs-Nijland, M. E.; Cassee, F. R.; Russell, A. G.; Brook, J. R.; Hopke, P. K.; Hidy, G.; Phalen, R. F.; Saldiva, P.; Sarnat, S. E.; Balmes, J. R.; Tager, I. B.; Ozkaynak, H.; Vedal, S.; Wierman, S. S. G.; Costa, D. L. Air Pollution and Health: Bridging the Gap from Sources to Health Outcomes: Conference Summary. Air Qual., Atmos. Health 2012, 5, 9–62. 95 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

3.

4.

5.

6.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

7.

8. 9. 10. 11.

12.

13.

14.

15.

16. 17.

18.

Uno, I.; Eguchi, K.; Yumimoto, K.; Takemura, T.; Shimizu, A.; Uematsu, M.; Liu, Z. Y.; Wang, Z. F.; Hara, Y.; Sugimoto, N. Asian Dust Transported One Full Circuit Around the Globe. Nat. Geosci. 2009, 2, 557–560. Fischer, E. V.; Hsu, N. C.; Jaffe, D. A.; Jeong, M. J.; Gong, S. L. A Decade of Dust: Asian Dust and Springtime Aerosol Load in the US Pacific Northwest. Geophys. Res. Lett. 2009, 36, 1–5. Capo, R. C.; Stewart, B. W.; Chadwick, O. A. Strontium Isotopes as Tracers of Ecosystem Processes: Theory and Methods. Geoderma 1998, 82, 197–225. Majestic, B. J.; Anbar, A. D.; Herckes, P. Stable Isotopes as a Tool to Apportion Atmospheric Iron. Environ. Sci. Technol. 2009, 43, 4327–4333. Hamelin, B.; Grousset, F. E.; Biscaye, P. E.; Zindler, A.; Prospero, J. M. Lead Isotopes in Trade-Wind Aerosols at Barbados - the Influence of European Emissions over the North-Atlantic. J. Geophys. Res.: Oceans 1989, 94, 16243–16250. Grousset, F. E.; Biscaye, P. E. Tracing Dust Sources and Transport Patterns Using Sr, Nd and Pb Isotopes. Chem. Geol. 2005, 222, 149–167. Isotope Tracers in Catchment Hydrology, 1st ed.; Kendall, C., McDonnell, J. J., Ed.;; Elsevier: New York, 1998; pp 51−86. Faure, G.; Mensing, T. M. Isotopes: Principles and Applications, 3rd ed.; Wiley: Hoboken; NJ, 2005; pp 75−106. Bollhofer, A.; Rosman, K. J. R. Isotopic Source Signatures for Atmospheric Lead: The Northern Hemisphere. Geochim. Cosmochim. Acta 2001, 65, 1727–1740. Veysseyre, A. M.; Bollhofer, A. F.; Rosman, K. J. R.; Ferrari, C. P.; Boutron, C. F. Tracing the Origin of Pollution in French Alpine Snow and Aerosols Using Lead Isotopic Ratios. Environ. Sci. Technol. 2001, 35, 4463–4469. Monna, F.; Lancelot, J.; Croudace, I. W.; Cundy, A. B.; Lewis, J. T. Pb Isotopic Composition of Airborne Particulate Material from France and the Southern United Kingdom: Implications for Pb Pollution Sources in Urban Areas. Environ. Sci. Technol. 1997, 31, 2277–2286. Simonetti, A.; Gariepy, C.; Carignan, J. Pb and Sr Isotopic Compositions of Snowpack from Quebec, Canada: Inferences on the Sources and Deposition Budgets of Atmospheric Heavy Metals. Geochim. Cosmochim. Acta 2000, 64, 5–20. Singer, A.; Zobeck, T.; Poberezsky, L.; Argaman, E. The PM10 and PM2.5 Dust Generation Potential of Soils/Sediments in the Southern Aral Sea Basin, Uzbekistan. J. Arid. Environ. 2003, 54, 705–728. Micklin, P. The Aral Sea disaster. Annu. Rev. Earth Planet. Sci. 2007, 35, 47–72. O’Hara, S. L.; Wiggs, G. F. S.; Mamedov, B.; Davidson, G.; Hubbard, R. B. Exposure to Airborne Dust Contaminated with Pesticide in the Aral Sea Region. Lancet 2000, 355, 627–628. Huang, X. T.; Oberhansli, H.; von Suchodoletz, H.; Sorrel, P. Dust Deposition in the Aral Sea: Implications for Changes in Atmospheric 96 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

19.

20.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

21.

22.

23.

24.

25. 26. 27.

28.

29.

30.

Circulation in Central Asia During the Past 2000 Years. Quat. Sci. Rev. 2011, 30, 3661–3674. Miller-Schulze, J. P.; Shafer, M. M.; Schauer, J. J.; Solomon, P. A.; Lantz, J.; Artamonova, M.; Chen, B.; Imashev, S.; Sverdlik, L.; Carmichael, G. R.; Deminter, J. T. Characteristics of Fine Particle Carbonaceous Aerosol at Two Remote Sites in Central Asia. Atmos. Environ. 2011, 45, 6955–6964. Dewan, N.; Majestic, B. J.; Ketterer, M. E.; Miller-Schulze, J. P.; Shafer, M. M.; Schauer, J. J.; Solomon, P. A.; Artamonova, M.; Chen, B. B.; Imashev, S. A.; Carmichael, G. Stable Isotopes of Lead and Strontium as Tracers of Sources of Airborne Particulate Matter in Kyrgyzstan. Atmos. Environ. 2015, 120, 438–446. Prospero, J. M.; Ginoux, P.; Torres, O.; Nicholson, S. E.; Gill, T. E. Environmental Characterization of Global Sources of Atmospheric Soil Dust Identified with the Nimbus 7 Total Ozone Mapping Spectrometer (TOMS) Absorbing Aerosol Product. Rev. Geophys. 2002, 40, 2-1–2-31. Mikova, J.; Denkova, P. Modified Chromatographic Separation Scheme for Sr and Nd Isotope Analysis in Geological Silicate Samples. J. Geosci. 2007, 52, 221–226. Chen, B. B.; Sverdlik, L. G.; Imashev, S. A.; Solomon, P. A.; Lantz, J.; Schauer, J. J.; Shafer, M. M.; Artamonova, M. S.; Carmichael, G. Empirical Relationship Between Particulate Matter and Aerosol Optical Depth Over Northern Tien-Shan, Central Asia. Air Qual., Atmos. Health 2013, 6, 385–396. Buat-Menard, P.; Chesselet, R. Variable Influence of the Atmospheric Flux on the Trace Metal Chemistry of Oceanic Suspended Matter. Earth Planet Sci. Lett. 1979, 42, 399–411. Taylor, S. R.; Mclennan, S. M. The Geochemical Evolution of the Continental-Crust. Rev. Geophys. 1995, 33, 241–265. Dickin, A. P. Radiogenic Isotope Geology, 2nd ed.; Cambridge University Press: Cambridge, U.K., 2005; pp 101−130 Monna, F.; Clauer, N.; Toulkeridis, T.; Lancelot, J. R. Influence of Anthropogenic Activity on the Lead Isotope Signature of Thau Lake Sediments (Southern France): Origin and Temporal Evolution. Appl. Geochem. 2000, 15, 1291–1305. Zhang, X. Y.; An, Z. S.; Liu, D. S.; Chen, T.; Zhang, G. Y.; Arimoto, R.; Zhu, G. H.; Wang, X. F. Study on 3 Dust Storms in China - Source Characterization of Atmospheric Trace-Element and Transport Process of Mineral Aerosol-Particles. Chin. Sci. Bull. 1992, 37, 940–945. Lin, Y. C.; Tsai, C. J.; Wu, Y. C.; Zhang, R.; Chi, K. H.; Huang, Y. T.; Lin, S. H.; Hsu, S. C. Characteristics of Trace Metals in Traffic-derived Particles in Hsuehshan Tunnel, Taiwan: Size Distribution, Potential Source, and Fingerprinting Metal Ratio. Atmos. Chem. Phys. 2015, 15, 4117–4130. Widory, D.; Liu, X. D.; Dong, S. P. Isotopes as Tracers of Sources of Lead and Strontium in Aerosols (TSP & PM2.5) in Beijing. Atmos. Environ. 2010, 44, 3679–3687.

97 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch004

31. Pettke, T.; Halliday, A. N.; Hall, C. M.; Rea, D. K. Dust production and deposition in Asia and the north Pacific Ocean over the past 12 Myr. Earth Planet. Sci. Lett. 2000, 178, 397–413. 32. Liu, C. Q.; Masuda, A.; Okada, A.; Yabuki, S.; Fan, Z. L. Isotope Geochemistry of Quaternary Deposits from the Arid Lands in Northern China. Earth Planet. Sci. Lett. 1994, 127, 25–38. 33. Chen, B. B.; Sverdlik, L.; Imashev, S.; Solomon, P. A.; Lantz, J.; Schauer, J. J.; Shafer, M. M.; Artamonova, A.; Carmichael, G. Lidar Measurements of the Vertical Distribution of Aerosol Optical and Physical Properties over Central Asia. Int. J. Atmos. Sci. 2013, 2013, 1–17. 34. Kulkarni, S.; Sobhani, N.; Miller-Schulze, J. P.; Shafer, M. M.; Schauer, J. J.; Solomon, P. A.; Saide, P. E.; Spak, S. N.; Cheng, Y. F.; Denier van der Gon, H. A. C.; Lu, Z.; Streets, D. G.; Jansens-Maenhout, G.; Wiedinmyer, C.; Lantz, J.; Artamonov, A.; Chen, B.; Imashev, S.; Sverdlik, L.; Deminter, J. T.; Adhikary, B.; D’Allura, A.; Wei, C.; Carmichael, G. Source Sector and Region Contributions to BC and PM2.5 in Central Asia. Atmos. Chem. Phys. 2014, 14, 11343–11392. 35. Komerek, M.; Ettler, V.; Chrastny, V.; Mihailovic, M. Lead Isotopes in Environmental Sciences: A Review. Environ. Int. 2008, 34, 562–577. 36. Schauer, J. J. Manuscript in preparation, 2015.

98 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Chapter 5

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

Analysis of Salts and Salt Substitutes with a Handheld X-Ray Fluorescence Analyzer Anthony Barakat, Shelby Maurice, Cameron Roberts, Mark A. Benvenuto, and Elizabeth S. Roberts-Kirchhoff* Department of Chemistry and Biochemistry, University of Detroit Mercy, 4001 W. McNichols Road, Detroit, Michigan 48221 *E-mail: [email protected].

Twenty-one consumer salts or salt substitutes purchased from local stores and three standard reference materials (SRMs) obtained from the National Institute of Standards and Technology (NIST) were analyzed using a handheld X-ray fluorescence (XRF) analyzer. Three of the commercial salts contained different colored grains, and these were further analyzed after separating the different colored particles. The XRF method allowed for the analysis of the elemental composition of a number of samples in under one hour. The results from the analysis of the SRMs with the handheld XRF were within 10.3% or less of the reported values for iron and strontium. Based on characteristic Kα and Kβ lines from the XRF spectral analysis of the salts, the salts contained potassium, iron, bromine, and strontium. A calibration curve for bromine was prepared by mixing known amounts of sodium bromide into known amounts of sodium chloride. The calibration curve was then used to determine the mass percent of bromine in the salts.

Introduction At one time, soldiers in the Roman Empire were paid in salt each day for their wages. The money they received was their “salarium” (from Latin) or salt-money. Today, salt is used as a commodity, to maintain health, as a food preservative, and even as a way to prevent crops from growing (1). The U.S. Geological Survey © 2015 American Chemical Society In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

maintains records on salt production in the world each year and reports on the production as open pan salt, solar salt, or rock salt. According to the 2015 Mineral Commodity Summary for salt, the estimated type of salt sold or used was rock salt (42%), salt in brine (40%), vacuum pan salt (10%), and solar salt (8%) (2). Most of the salt mined or recovered in the world is used in highway deicing and as chemical feedstock to produce sodium hydroxide and chlorine gas. Only a small percentage (4%) of the salt mined or produced is used in the food industry (1). Salt is primarily obtained from sea water or mined as rock salt or halite. The consumer salts used in the food industry are isolated, refined and ground. Many times an anticaking agent is also added. Consumers can purchase their table salt as sodium chloride with or without added iodine. Iodine is added to prevent iodine deficiency since iodine is needed to synthesize hormones produced by the thyroid gland including thyroxin and triiodothyronine. Consumers can also purchase various sea salts that are isolated by the solar or open pan method. The approximate composition by mass of the major ions in sea water include 56% chloride, 31% sodium, 4% magnesium, 8% sulfate, 1% calcium, and 1% potassium (3). Depending on where the sea salt is obtained, it can vary in taste from other salts. This variety is due to the environment from which it was obtained including clays, algae, and plants in the local waters. The components other than sodium chloride can affect the color and taste of the salt. In addition, the size of the grain can have an impact on the interaction on the tongue and thus the taste of the salt (4). Given the differences in taste of these sea salts, there is evidence that there may be differences in the ability to flavor dishes, but the evidence that they are better for your health is not as strong. In addition to sodium chloride, salt substitutes are used to flavor foods. These are mainly potassium chloride and are used to alleviate the health problems including high blood pressure and cardiovascular disease attributed with a high-salt (sodium chloride) diet (5). X-ray fluorescence (XRF) analysis has been used with various food and food supplements including kelp and clay supplements, vitamins, calcium in milk powder, bromine in flours and bakery products, potassium and sodium in marine salts, food premixes, FD&C (Food, Drug, & Cosmetics) dyes, teas, soft drinks, fruit juices, infant cereals, Ayurvedic medicines, and fast foods (6–30). In XRF spectroscopy, radiation from an X-ray source ejects electrons from an atom’s first (K) and second (L) inner shells. Electrons from higher shells then fill the empty shells. When these electrons enter the lower state, characteristic X-ray fluorescence is emitted. When an electron from the K shell is ejected by an external X-ray and a L electron takes its place, a Kα photon is emitted. This results in the Kα line on the spectrum. When an electron from the K shell is ejected by an external X-ray and a M electron takes its place, a Kβ photon is emitted. This results in the Kβ line on the spectrum. There are unique emission spectra for elements with atomic numbers 11-92. The strength of the emission from these can be used to quantitate the amount of an element (31–33). Portable or handheld XRF analyzers can be used to rapidly scan a variety of samples for a number of different elements. In this study, twenty-one commercially available salts or salt substitutes were analyzed with a Bruker S1 Turbo handheld XRF analyzer to compare their elemental compositions. 102 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

Experimental Methods Standard reference materials (SRMs) were obtained from the National Institute of Standards and Technology (NIST). SRM 2586 is the Trace Elements in Soil standard containing lead from paint (Nominal 500 mg lead/kg) (34). SRM 2587 is the Trace Elements in Soil standard containing lead from paint (Nominal 3000 mg lead/kg) standard (35). Both of these soil standards have reported certified mass fractions for arsenic, cadmium, chromium, and lead. These SRMs were used as standards for these other elements and not for the quantity of lead present. They also have reported reference mass fractions for many elements including calcium, iron, manganese, potassium, titanium, and zinc (34, 35). SRM 2709a is the San Joaquin Soil standard. It has certified mass fractions for calcium, potassium, lead and strontium (36). A Bruker S1 TURBO handheld XRF analyzer with a silicon drift detector and a resolution of approximately 145 eV at 200000 cps was used to analyze the samples and SRMs as previously described (6). For each SRM, three separate samples (3.0 g per sample) were analyzed five times each for 60 s. The instrument was mounted on the bench-top stand. Each sample was placed on the safety platform to ensure the samples were analyzed from an equal distance. The instrument used a voltage of 45 kV and a current of 30 µA, with a Ti/Al filter. A Hewlett Packard PDA with XBruker Elemental S1 software was used to control the instrument, and data were collected and analyzed with the soil fundamental parameter (FP) calibration method. This calibration has been optimized for a SiO2 matrix, is non-normalized, and relies on using a repeatable distance between the sample and detector. The reported elements with this method include K, Ca, Ti, V, Cr, Mn, Fe, Co, Ni, Cu, Zn, As, Se, Br, Rb, Sr, Zr, Nb, Mo, Pd, Ag, Cd, Sn, Sb, Ba, Ce, Hf, Ta, W, Pt, Au, Hg, Tl, Pb, Bi, Th, and U. The peak height intensities for each peak in counts per second (cps) and statistical analyses were determined after importing the data file into Microsoft Excel as described (25). Characteristic peaks for each element were identified using S1PXRF software from Bruker. The SRMs, used as provided, were placed into Chemplex Spectrocertified Quality XRF micro-sample cups (31.0 mm × 22.4 mm) and were covered with Chemplex Spectromembrane perforated thin film mylar polyester sample support carrier films (3.6 µm) with a sealing ring. The concentrations (averages and standard deviations) of various elements for each of the standards were determined and compared to the NIST standard values (34–36). The limit of detection (LOD) and limit of quantitation (LOQ) for each element in soil analyzed with the soil FP calibration were provided by Bruker (37). The values of interest for this study are shown in Table 1. The LOQ was defined as five times the LOD for this specific method. Bromine standards were produced using high purity sodium bromide (≥99%) in high purity sodium chloride (≥99.0%). These reagents were from Sigma-Aldrich. The samples were ground with a mortar and pestle for 30 s and then homogenized in a dedicated coffee grinder prior to transfer to sample containers (3.0 g per sample). Twenty-one salts or salt substitutes were analyzed with a Bruker S1 TURBO handheld XRF instrument as described for the SRMs. Each of the salt samples were ground using a mortar and pestle for 30 s and measured out to five equal samples (3.0 g per sample) and analyzed for 60 s. 103 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Table 1. Limits of Detection and Quantitation for Selected Elements with the S1 Turbo Handheld XRF Analyzer Using the Soil FP Calibration Method (37)

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

a

K

Ca

Ti

Mn

Fe

Br

Sr

Zn

Pb

LODa (ppm)

450

200

90

18

17

1

3

3

7

LOQb (ppm)

2250

1000

450

90

85

5

15

15

35

LOD, Limit of Detection.

b

LOQ, Limit of Quantitation.

Results and Discussion The NIST standards were analyzed with the soil calibration method with the handheld XRF analyzer. The concentrations in ppm or mg/kg (averages ± standard deviations) for potassium, strontium, and bromine in the three NIST standards are shown in Table 2. The percent relative standard deviations are also given. The averages were compared to the reference values for SRMs 2586 and 2587 and certified values for SRM 2709a and expressed as percent error. SRM 2586 is the Trace Elements in Soil standard containing lead from paint (Nominal 500 mg lead/kg) (34). SRM 2587 is the Trace Elements in Soil standard containing lead from paint (Nominal 3000 mg lead/kg) standard (35). SRM 2709a is the San Joaquin Soil standard (36). Results from the elements of interest for this study are presented. The percent errors for the values for potassium are quite high. With XRF analysis, the errors are larger for the lighter elements. The yields for these elements are lower for the instrument since the emitted X-rays are at low energies (33). The percent error for iron is less than 10% in all three standards. Another element of interest in this study is strontium and the percent error for this is 10.3% or less for each standard. The relative standard deviation for each is given in the parentheses and is less than 9.5%.

Consumer Salts Twenty-one different consumer salts or salt substitutes were purchased from various stores and analyzed using the soil calibration method with the handheld XRF analyzer. This method was selected since it reports results from 37 elements. Each salt was given a sample identification designation (Table 3). The salt designated sb is the common Morton’s Iodized salt. The salts s1-s3 are salt substitutes recommended for those who require a lower-salt diet. The others are various kosher and sea salts. Some of these salts have significant color variations in the salt grains. For example, s12 and s13 have a mixture of red and white grains, while s17 has a mixture of blue and white grains. 104 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Table 2. Concentrations of Potassium, Iron, and Strontium from Analysis with the Handheld XRF Compared to the NIST Reported Values for the SRMsa

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

K (ppm) 2586 XRFc

6781 ± 307 (4.5)d

2586 SRM

9760 ± 180 (1.8)

2587 XRF

11420 ± 784 (6.9)

2587 SRM

15830 ± 550 (3.5)

2709a XRF

12525 ± 469 (3.8)

2709a SRM

21100 ± 600 (2.8)

% errorb

30.5

27.8

40.6

Fe (ppm) 48200 ± 1230 (2.6) 51610 ± 890 (1.7) 27090 ± 1665 (6.2) 28130 ± 250 (0.9) 30408 ± 1062 (6.2) 33600 ± 700 (1.9)

% error

6.6

3.8

9.5

Sr (ppm) 75.4 ± 3.2 (4.2) 84.1 ± 8.0 (9.5) 128.0 ± 2.3 (1.8) 126 ± 19 (1.8) 230.3 ± 4.5 (2.0) 239 ± 6 (1.8)

% error

10.3

1.6

3.6

All values for SRM 2587 and 2586 are reference values (34, 35). Values for SRM 2709a are certified values (36). b Percent error between measured XRF value and NIST reported value. c Values (average ± standard deviation) for XRF from analysis with handheld instrument. d Percent relative standard deviation. a

The results for potassium are shown in Figure 1, and the results for iron, bromine, and strontium are shown in Figure 2. These four elements were identified at various levels in some of the different salts. In the samples, these were the only elements detected over the LOD of the instrument. The presence of these elements was verified by looking at the spectra and identifying the characteristic Kα and Kβ peaks. Potassium was identified with a Kα peak at 3.31 keV and Kβ peak at 3.58 keV from a representative spectrum. As expected, significant amounts of potassium were found in the salt substitutes, s1, s2, and s3, since potassium chloride is used as a substitute for sodium chloride. In addition, levels of potassium over the LOQ of 2250 ppm were also found in s8, Stonewall Kitchen Maine Sea Salt, and s17, All Natural Blue Persian Salt Mill. Figure 2 shows the results from the analysis of iron, bromine, and strontium in the various salt samples. For iron, the LOQ in soil is 75 ppm. A representative spectrum for s17 showed the Fe Kα peak at 6.36 keV and Kβ peak at 7.06 keV. Seven of the salt samples had Fe concentrations greater than this LOQ. These are s4, 12, 13, 15, 18, 19 and 20. For strontium, the LOQ is 15 ppm. The salts containing strontium were identified using the characteristic emission peaks. For example, a representative spectrum for s17 contained a strontium Kα peak at 14.08 keV and Kβ peak at 15.79 keV. Many of the salts had a concentration of strontium over the LOQ including salts s6-10, s12, 105 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

s13, s15, s17-20. These are all sea salts and strontium has been identified in sea water (38). The highest amount was found in s20, 136.2 ppm. For bromine the LOQ is 5 ppm. A representative spectrum for s17 showed a bromine Kα peak at 11.86 keV and Kβ peak at 13.21 keV. The only salts that did not have at least this amount of bromine were s1, s3, and s4. The samples s8 and s17 had the highest concentrations of bromine. Bromine is often isolated from bromine-rich sea brines (39).

Table 3. Consumer Salts and Salt Substitutes Analyzed by XRF Designation

Salt

sb

Morton Iodized Salt

s1

Morton Lite Salt

s2

The Original Lo Salt

s3

The Original No Salt

s4

Redmond Real Salt

s5

David’s Kosher Salt

s6

Le Paludier Fleur de Sel

s7

Le Saunier De Camargue Fleur De Sel

s8

Stonewall Kitchen Maine Sea Salt

s9

Eden Sea Salt

s10

Roland Sea Salt

s11

La Baleine Sea Salt

s12

Olde Thompson Himalayan Pink Salt

s13

Cerulean Seas Pink Himalayan

s14

Himala Salt

s15

Celtic Sea Salt

s16

All Natural Mediterranean Salt Mill

s17

All Natural Blue Persian Salt Mill

s18

All Natural Grey Brittany Sea Salt Mill

s19

All Natural Pink Himalayan Salt Mill

s20

Beyond the Shaker Alderwood Smoke Salt

s21

Old Time Hawaiian Sea Salt

106 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

Figure 1. Concentration of potassium in the different salts as determined by XRF analysis.

Figure 2. Concentration of iron, bromine, and strontium in salt samples by XRF analysis. Using the soil calibration method with these samples is a convenient way to compare the elemental composition in these salts but is not the best method to determine the concentration of each element since the salts have a different matrix from soil. The amount of bromine varied in the salts so it was of interest to further investigate this with a more quantitative study. To determine the mass fractions of bromine in the various salt samples, a standard curve was generated using high 107 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

purity NaBr in high purity NaCl. The samples included a blank and four bromine standards in the NaCl matrix. A portion of the representative spectra from the blank and four standards are shown in Figure 3. The intensity (peak height) of the Kα peak at 11.84 ± 0.02 keV, characteristic for bromine, was used to generate the standard curve shown in Figure 4.

Figure 3. A portion of the spectra with the characteristic Br K peaks for the blank and four concentrations (mass fractions) of bromine in sodium chloride.

Figure 4. Standard curve of the mass fraction of bromine. The results are given as average ± standard deviation for each concentration. The RSDs ranged from 3.7-6.8% and the R2 value was 0.9905 showing the linear relationship between the amount of bromine and intensity of the Kα peak. 108 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

The Br calibration curve was used to determine the mass fraction of bromine in each of the salts. The spectra from the salts were analyzed and the intensities of the bromine Kα peaks were determined for each. This was then converted to the mass fraction using the calibration curve. The average mass fraction Br (%) for each of the salts is shown in Figure 5. The salts s8 and s17 have the highest concentrations of Br. These high levels of bromine are found in the two NaCl salts that had the highest concentrations of potassium (Figure 1). Stonewall Kitchen Main Sea Salt (s8) is isolated from Maine and All Natural Persian Salt Mill (s17) is isolated from Iran. These similarities are not a result of isolation from the same area but from areas that have similar elements in the environments (4).

Figure 5. Average mass fraction of Br in each of the salts.

Some of the salts have grains of different colors. For example s12, Olde Thompson Himalayan Pink Salt, has both red and white grains. These were manually separated into the red and white pieces as shown in Figure 6. This separation was also done with s13, which has red and white grains, and s17, which has blue and white grains. The different colored grains were then ground and analyzed separately. Table 4 shows the results looking at the presence or absence of bromine and strontium in three salts, s12, s13, and s17. All of the salt samples contained bromine when analyzed by XRF. Strontium was present in the red grains in s12 and s13 and in the white grains in s17. The spectra for s17 without separation, and the blue and white grains are shown in Figure 7. The bromine peaks are in both the blue and white grains while the Sr Kα peak is found in the unseparated mixture and the white grains. When the colored grains were separated, the results from the XRF analysis showed that the different colored grains from the same salt did not have the same composition. Also the dark colored grains from s12, s13, and s17 had different colors (red and blue) and compositions since strontium was found in the dark grains in s12 and s13 and in the white grains in s17. Further analysis may lead to determining if there is a correlation between the darker colors and the presence of other elements. 109 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

Figure 6. The red and white portions of s12, Olde Thompson Himalayan Pink Salt. Credit Elizabeth S. Roberts-Kirchhoff.

Table 4. Presence of Bromine and Strontium after Separation of the Colors in the Salt Salt Olde Thompson Himalayan Pink Salt

Cerulean Seas Pink Himalayan

All Natural Blue Persian Salt

Salt number and color

Br

Sr

s12 both

+

+

s12 red

+

+

s12 white

+

-

s13 both

+

+

s13 red

+

+

s13 white

+

-

s17 both

+

+

s17 blue

+

-

s17 white

+

+

110 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

Figure 7. XRF spectrum of s17 without separation, s17 white grains, and s17 blue grains.

Conclusions A handheld XRF using a soil calibration method was used to analyze three soil standard reference materials and twenty-one consumer salts or salt substitutes. The results from the soil method can be used to quickly compare the different salts for the different elements monitored in this method. The spectra were used to verify that an element was present by looking at the characteristic K peaks. Potassium, iron, bromine, and strontium were identified in some of the salts. In addition, the intensity of characteristic peaks was determined from the spectra. A calibration curve for bromine in a sodium chloride matrix was generated. This was then used to quantitate the mass fraction of Br in the samples. In addition, when the colored grains were separated in three of the salts, the bromine was presented in both colors but the strontium was present in only one of the colors. This method is a quick way to analyze a large number of samples. An instrument-supplied calibration method such as the soil calibration can be used to compare between samples and a calibration curve can be generated using a more similar matrix in order to quantitate the concentrations of an element in a unique matrix.

Acknowledgments The University of Detroit Mercy’s McNichols Faculty Assembly Internal Research Fund and the Michigan-Ohio University Transportation Center Grant provided funding. 111 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

References 1. 2.

3. 4.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

5.

6.

7.

8. 9. 10.

11.

12.

13.

14.

15.

Benvenuto, M. A. Industrial Chemistry; Walter de Gruyter GmbH: Berlin, 2014; pp 31−33. U.S. Department of the Interior. Mineral Commodity Summaries 2015, U.S. Geologic Survey; http://minerals.usgs.gov/minerals/pubs/mcs/ (accessed July 2015). Great Salt Lake Utah Geological Survey; http://geology.utah.gov/popular/ general-geology/great-salt-lake/ (accessed July 2015). Drake, S. L.; Drake, M. A. Comparison of Salty Taste and Time Intensity of Sea and Land Salts from Around the World. J. Sens. Stud. 2011, 26, 25–34. Strategies to Reduce Sodium Intake in the United States; Henney, J. E., Taylor, C. T., Boon, C. S., Eds.; Institute of Medicine of the National Academies, The National Academies Press: Washington, DC, 2010. Labond, J. L.; Stroeters, N. H.; Benvenuto, M. A.; Roberts-Kirchhoff, E. S. Analysis of Nine Edible Clay Supplements with a Handheld XRF Analyzer. In Chemistry of Food and Food Contact Materials: From Production to Plate; Benvenuto, M. A., Ahuja, S., Duncan, T. V., Noonan, G. O., RobertsKirchhoff, E. S., Eds.; ACS Symposium Series 1159; American Chemical Society: Washington, DC, 2014; pp 99−111. Shroeder, A. E.; Smith, Z. R.; Benvenuto, M. A.; Roberts-Kirchhoff, E. S. Analysis of Dietary Supplements with a Handheld XRF Analyzer. In Food, Energy, and Water: The Chemistry Connection; Ahuja, S., Ed.; Elsevier, Inc: Amsterdam, Netherlands, 2015; pp 383−390. Anderson, D. L. Determination of bromine in regulated foods with a field portable X-ray fluorescence analyzer. J. AOAC Int. 2009, 92, 502–510. Anderson, D. L. Analysis of beverages for Hg, As, Pb, and Cd with a field portable X-ray fluorescence analyzer. J. AOAC Int. 2010, 93, 683–693. Anderson, D. L. Screening of foods and related products for toxic elements with a portable X-ray tube analyzer. J. Radioanal. Nucl. Chem. 2009, 282, 415–418. Palmer, P.; Jacobs, R.; Baker, P. E.; Ferguson, K.; Webber, S. Use of FieldPortable XRF Analyzers for Rapid Screening of Toxic Elements in FDARegulated Products. J. Agric. Food Chem. 2009, 57, 2605–2613. Perring, L.; Audrey, D.; Basic-Dvorzak, M.; Hammer, D. Rapid quantification of iron, copper, and zinc in food premixes using energydispersive XRF. J. Food Compos. Anal. 2005, 18, 655–663. Nielson, K. K.; Mahoney, A. W.; Williams, L. S.; Rogers, V. C. X-ray fluorescence measurements of Mg, P, Cl, K, Ca, Mn, Fe, Cu, and Zn in fruits, vegetables, and grain products. J. Food Compos. Anal. 1991, 4, 39–51. Perring, L.; Andrey, D.; Basic-Dvorzak, M.; Blanc, J. Rapid Multimineral Determination in Infant Cereal Matrices Using Wavelength Dispersive X-ray Fluorescence. J. Agric. Food Chem. 2005, 53, 4696–4700. Nielson, K. B.; Mahoney, A. W.; Williams, L. S.; Rogers, V. C. Mineral Concentrations and Variations in Fast-Food Samples Analyzed by X-ray Fluorescence. J. Agric. Food Chem. 1991, 39, 887–892. 112 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

16. Mohapatra, A.; Rautray, T. R.; Vijayan, V.; Mohanty, R. K.; Dey, S. K. Trace elemental characterization of some food crustacean tissue samples by EDXRF technique. Aquaculture 2007, 270, 552–558. 17. Da-Col, J. A.; Bueno, M. I. M.; Melquiades, F. L. Fast and Direct Na and K Determination in Table, Marine, and Low Sodium Salts by X-ray Fluorescence and Chemometrics. J. Agric Food Chem 2015, 63, 2406–2412. 18. Perring, L.; Audrey, D. EDXRF as a Tool for Rapid Minerals Control in Milk-Based Products. J. Agric. Food Chem. 2003, 51, 4207–4212. 19. Verbi Pereira, F. M.; Pereira-Filho, E. R.; Rodriques, E.; Maretti Silveira Bueno, M. I. Development of a methodology for Ca, Fe, K, Mg, Mn, and Zn quantification in teas using X-ray spectroscopy and multivariate calibration. J. Agric. Food Chem. 2006, 54, 5723–5730. 20. Bao, S. X.; Wong, Z. H.; Liu, J. S. X-ray fluorescence analysis of trace elements in fruit juice. Spectrochim. Acta B 1999, 54, 1893–1897. 21. Zucchi, O. L.; Moreira, S.; Salvador, M. J.; Santos, L. L. Multi-element analysis of soft drinks by X-ray fluorescence spectrometry. J. Agric. Food Chem. 2005, 53, 7863–7869. 22. Hepp, N. M. Spectroscopic determination of chromium in FD&C Blue No. 1 by X-ray fluorescence. J. AOAC Int. 1996, 79, 1189–1190. 23. Saper, R. B.; Kales, S. N.; Paquin, J.; Burns, M. T.; Eisenberg, D. M.; Davis, R. B.; Phillips, R. S. Heavy metal content of Ayurvedic herbal medicine products. J. Am. Med. Assoc. 2004, 292, 2868–2873. 24. Garshott, D. M.; Macdonald, E. A.; Murray, M. N.; Benvenuto, M. A.; Roberts-Kirchhoff, E. S. Elemental Analysis of a Variety of Dried, Powdered, Kelp Food Supplements for the Presence of Heavy Metals via Energy-Dispersive X-Ray Fluorescence Spectrometry. In It’s All In the Water: Studies of Materials and Conditions in Fresh and Salt Water Bodies; Benvenuto, M. A., Roberts-Kirchhoff, E. S., Murray, M. N., Garshott, D. M., Eds.; ACS Symposium Series 1086; American Chemical Society: Washington, DC, 2011; pp 123−133. 25. Chan, J. C.; Palmer, P. T. Determination of calcium in powdered milk via X-ray fluorescence using external standard and standard addition based methods. J. Chem. Educ. 2013, 90, 1218–1221. 26. Gao, H.; Zhao, H.; Wang, Z.; Zhou, J.; Guan, Y.; Shi, J. Analysis on spirulinas by PXRD and XRF. Shipin Keji 2008, 12, 270–272. 27. Pajchel, L.; Nykiel, P.; Kolodziejski, W. Elemental and structural analysis of silicon forms in herbal drugs using silicon-29 MAS NMR and WD-XRF spectroscopic methods. J. Pharm. Biomed. Anal. 2011, 56, 846–850. 28. Saleh-e-in, M. M.; Sultana, A.; Hossain, M. A.; Ahsan, M.; Roy, S. K. Macro and micro elemental analysis of Anethum sowa L. (Dill) stem by X-ray fluorescence spectrometry. Bangladesh J. Sci. Ind. Res. 2008, 43, 479–490. 29. Janosova, V.; Sykorova, M.; Stroffekova, O.; Havranek, E. Determination of selected elements in dietary supplements by X-ray fluorescence analysis. Farm. Obz. 2008, 77, 177–181.

113 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch005

30. Raggi, M. A.; Schiavone, P.; Lucchini, F.; Mandrioli, R. Determination of selenium in diet supplement by spectrometry and X-ray fluorescence. Boll. Chim. Farm. 1995, 134, 574–579. 31. XRF Research, Inc.; X-Ray Fluorescence; http://www.xrfresearch.com/ technology/x-ray-fluorescence (accessed April 2015). 32. Geochemical Instrumentation and Analysis. X-ray Fluorescence Instrumentation; http://serc.carleton.edu/research_education/ geochemsheets/techniques/XRF.html (accessed April 2015). 33. Oxford Labs. X-Ray Fluorescence: Qualitative and Quantitative Analysis; http://oxford-labs.com/x-ray-fluorescence/qualitative-and-quantitativeanalysis/ (accessed July 2015). 34. Wise, S. A.; Watters, R. L. Certificate of Analysis SRM® 2586 Trace Elements in Soil Containing Lead from Paint (Nominal 3000 mg/kg Lead); National Institute of Standards & Technology: Gaithersburg, MD, 2008. 35. Wise, S. A.; Watters, R. L. Certificate of Analysis SRM® 2587 Trace Elements in Soil Containing Lead from Paint (Nominal 3000 mg/kg Lead); National Institute of Standards & Technology: Gaithersburg, MD, 2008. 36. Wise, S. A.; Watters, R. L. Certificate of Analysis SRM® 2709a San Joaquin Soil; National Institute of Standards & Technology: Gaithersburg, MD, 2009. 37. Bruker Corporation. S1 Turbosdr Calibrations; Bruker Elemental: Kennewick, WA. November 2010. 38. Culkin, F.; Cox, R. A. Sodium, potassium, magnesium, calcium, and strontium in sea water. Deep Sea Research and Oceanographic Abstracts. 1976, 13, 789–804. 39. Emsley, J. Bromine. In Natures’s Building Blocks: An A-Z Guide to the Elements; Oxford University Press: Oxford, 2001; pp 69−73.

114 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Chapter 6

Mercury-Thiourea Complex Ion Chromatography: Advances in System Chemistry and Applications to Environmental Mercury Speciation Analysis Todd A. Olsen,1,4 Tina H. Huang,2 Ramdas Kanissery,3 and Robert J. M. Hudson*,3 1Department

of Civil and Environmental Engineering, University of Illinois at Urbana-Champaign, Urbana, Illinois 61801 2Department of Chemistry, University of Illinois at Urbana-Champaign, 367B Noyes Lab, 601 S. Mathews Avenue, Urbana, Illinois 61801 3Department of Natural Resources and Environmental Sciences, University of Illinois at Urbana-Champaign, W-503 Turner Hall, 1102 S. Goodwin Avenue, Urbana, Illinois 61801 4Current address: Division of Environmental Sciences, Oak Ridge National Laboratory, P.O. BOX 2008 MS 6036, Oak Ridge, Tennessee 37931 *E-mail: [email protected].

Mercury-thiourea complex ion chromatography is the core of a relatively new approach to Hg speciation analysis that is sensitive enough to accurately quantitate monomethyl Hg at ultratrace environmental levels. In this chapter, a detailed description of an updated system chemistry and operating conditions for performing mercury speciation analysis using inductively coupled plasma mass spectrometry are presented. The new operating conditions are very stable, highly-sensitive, and free of problems that afflicted earlier versions of the system. The potential for obtaining accurate results using the system is demonstrated by the excellent performance metrics obtained and the close agreement of these results with consensus values of monomethyl Hg in water and sediment reference materials and inter-lab comparison samples. The ability to quantitate © 2015 American Chemical Society In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

mercuric and monoethyl along with monomethyl forms of Hg is a feature of this approach that likely will prove useful in future studies of environmental systems.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Environmental Hg Speciation Despite the analytical challenges of accurately measuring multiple chemical forms of mercury at ultratrace levels, environmental chemists, hydrologists, and oceanographers have made much progress in understanding the distribution and biogeochemical cycling of this element in aquatic ecosystems over the past 25-30 years. The central fact established by their work is that the levels of Hg are very low, except in systems receiving direct discharges of Hg pollution. In most of the atmosphere, the partial pressure of elemental mercury is ~10–13 atm (1). In marine and fresh surface waters, dissolved Hg occurs at nanogram per liter (picomolar) levels or less (2). In sediments and soils, Hg levels are roughly 1000-fold higher on a mass basis but still fall in the tens of nanograms per gram range (3). In biota, levels commonly reach into the hundreds of nanograms per gram, making tissues the easiest to analyze of the main environmental media (4). The fact that Hg naturally occurs at such low levels not only makes chemical analysis more difficult, but it explains why humans have been able to significantly perturb Hg levels in environmental systems ranging from local to global in scale (5). Environmental chemists have also shown that most Hg in the environment occurs in a few distinct chemical forms – elemental, mercuric, monomethyl and dimethyl – with the proportions of each varying between environmental compartments (6, 7). Their investigations of Hg speciation have employed a wide variety of analytical methods in different environmental media. Our goal in this chapter is not to describe them all but to report on the potential of one new approach to Hg analysis to expand our understanding of environmental Hg speciation in aquatic ecosystems. To see this potential clearly, it is helpful to appreciate what has and has not been learned by employing the most widely-used methods. Major Forms of Dissolved Hg Environmental chemists seeking to fully characterize the speciation of dissolved mercury in natural waters generally divide the total (THg) among four major chemical forms: elemental (Hg0), mercuric (HgII), monomethyl (MMHg), and dimethyl (DMHg) (Figure 1). Using widely-known analytical methods, ambient concentrations of all forms except HgII are directly measureable. Monoethyl mercury (ETHg) has been reported only rarely and is generally neglected. Unlike most metals, Hg is “atmophilic.” As a result, not only do volatile forms of Hg occur in the atmosphere, but they also dissolve into natural waters. In freshwater systems, the only dissolved gaseous species typically present is Hg0. Thus, it can be measured as dissolved gaseous mercury (DGM) without performing the chromatographic step needed to distinguish DMHg and Hg0 in 116 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

marine waters. However, since Hg0 is not directly relevant to assessing the toxicity of Hg pollution, DGM is generally not measured in freshwater samples. Furthermore, since Hg0 oxidizes to HgII during sample storage, measurements of either one in such samples cannot be assumed to reflect ambient conditions. A further complication in Hg speciation analysis arises from the fact that measurements of “dissolved” species in natural waters are by necessity operationally-defined, i.e., by filtration using a filter with pores of diameter 0.4-µm or some similar size. Thus, it comes as no surprise that some chemists also report detecting nanoparticulate Hg capable of passing through these pores (Figure 1). Methods of quantifying HgNP involve isolation of the fine particulate fraction by ultrafiltration or solid-phase extraction, since at least some of the nanoparticles are hydrophobic. It should be noted that this Hg fraction is not often measured but can be a significant part of the total “dissolved” pool (8, 9).

Figure 1. Chemical forms of dissolved mercury in natural waters: nanoparticulate (HgNP), mercuric (HgII), elemental (Hg0), dimethyl (DMHg), monomethyl (MMHg), and monoethyl (ETHg). Aggregate analytes: Total (THg) and Dissolved Gaseous Mercury (DGM). Inferred: Inorganic (InHg). Directly-measurable forms in red italics.

The complete speciation scheme of Figure 1 is impractical for widespread use in environmental monitoring. Instead, the current practice in environmental assessment work is to monitor both MMHg, due to its great toxicological relevance, and THg. Datasets comprising these measures often derive a concentration of “inorganic Hg” (InHg), which includes HgII plus HgNP and any Hg0 that was originally present in the sample, from the difference between THg and MMHg:

Note that this nomenclature (Figure 1; Table 1) follows widely-adopted conventions of environmental mercury chemists. These parameters refer to measurable quantities of different analytes, which in several cases comprise groups of species, e.g., all species containing the element Hg or all complexes of monomethyl mercury. The same notation is also used when referring to the analytical method. 117 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Table 1. Nomenclature Used in Defining Forms of Hg Quantifiable by Analysis as Well as Distinct Chemical Species

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Analytical parameters THg

Total mercury (all species of Hg)

HgII

Mercuric mercury (except nanoparticulate)

HgI

Mercurous mercury

MMHg

Monomethyl mercury (all complexes of MeHg+)

DMHg

Dimethyl mercury (Me2Hg)

ETHg

Monoethyl mercury (all complexes of EtHg+)

InHg

Inorganic Hg (all except MMHg, DMHg, and ETHg)

DGM

Dissolved gaseous mercury (Me2Hg and Hg0)

HgNP

Nanoparticulate Hg (including HgS(s))

Distinct species (complexes not shown) Hg0

Elemental mercury

Hg2+

Mercuric ion

MeHg+

Monomethyl mercuric ion

Me2Hg

Dimethyl mercury

EtHg+

Monoethyl mercuric ion

MeHgEt

Methylethyl mercury

Hg22+

Mercurous ion

Hg-binding ligands considered OH–

Hydroxide ion

Cl–

Chloride ion

Br–

Bromide ion

SH–

Bisulfide ion

TU

Thiourea

HSRAA

Thiol-containing amino acids

HSRDOM

Thiol moieties of natural dissolved organic matter

These symbols and acronyms, however, are not necessarily the clearest for describing individual ionic or complex species and conflict with other naming conventions, such as “Me” for methyl. Thus, herein we employ a second convention – Hg2+, MeHg+, and EtHg+ – when referring to distinct species or the common central ion of a group of complexes. Complexes of these ions with 118 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

relevant ligands (Table 1) are written as normal central ion-ligand combinations. Finally, we also refer to groups of complexes with the same central metal ion and ligand, but variable ligand numbers as , where zL indicates the appropriate net charge of the complexes containing metal ion M and x molecules of ligand L.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Complexation of Dissolved Hg One central goal of environmental chemists is to characterize the speciation of each “truly dissolved” form of Hg (Figure 2). While knowing the concentrations of dissolved HgII and MMHg is enough to quantify their respective transport fluxes, knowing the concentrations of the aquo Hg2+ and MeHg+ ions and their coordination complexes in different environmental compartments is crucial to predicting their reactivities (10). The distribution of Hg2+ and MeHg+ among their various weak and strong complexes affects their bioavailabilities as well. The former include complexes with hydroxide (OH–) and chloride (Cl–) anions while the latter include complexes with bisulfide ion (SH–) and the thiol moieties of amino acids (HSRAA) and natural dissolved organic matter or DOM (HSRDOM).

Figure 2. Important complexes of Hg2+ and MeHg+ included in HgII and MMHg, along with Hg in nanoparticles. Arrows represent relatively rapid (solid) and slow (dashed) reactions. Free ligands not depicted. (see color insert) 119 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Methods for characterizing the organic complexation of Hg2+ are not as developed as they are for other environmentally-significant divalent metal ions, such as Cu2+. The few measurements that exist suggest that this ion, as well as MeHg+, are strongly bound (11–13) and it is widely accepted that this is more often the case than not (14). Note that essentially all measurements of Hg2+ complexation by natural ligands are limited by the operational nature of the measures of HgII employed. Those using THg measurements include both MMHg and HgNP in HgII when they should not, and those using SnCl2-reducible Hg may underestimate HgII. In fact, strong complexation by natural organic matter must also routinely be overcome in the sample preparations used in quantifying dissolved THg and MMHg in natural waters.

Major Forms of Hg in Sediments Saturated sediments are especially important compartments within aquatic ecosystems for Hg biogeochemistry because the anaerobic conditions that commonly form within them support the methylation of mercury (15). As these conditions typically do not develop in surface waters, sediments are usually the main source of MeHg+ to aquatic ecosystems. Thus, measuring the concentrations of the substrate for the methylation process – HgII – and the product – MMHg – both in bulk sediment samples and porewater has proven to be important in environmental work. Now many of the same Hg species and complexation reactions occur in sediment porewater as in surface waters, although adsorption reactions take on much greater significance. Adsorption is essentially a class of metal complexation reactions by the ligands located on the surfaces of particles (16). Thus, adsorption of both MeHg+ and Hg2+ by thiols in sediment organic matter and by sulfide minerals play a very significant role (10, 17). For Hg2+, formation of pure mineral phases, including cinnabar (α-HgS) and metacinnabar (β-HgS), can also limit porewater concentrations of HgII (18–20). While the use of direct solid phase speciation methods for Hg is increasing (21), much analysis of aquatic sediments still relies on chemical digestions to extract the analytes from the complex sample matrices (22). The initial stages of these sediment sample preparations usually differ markedly from those for water; by the final steps where the different forms of Hg are quantified, the methods are often related. In addition, one important consideration in designing methods is similar: how can any particular Hg species (the analyte) be extracted from an environmental matrix that strongly binds it? Thus, in this chapter, we also consider the analysis of MMHg in sediments.

Hg Speciation Analysis The understanding of Hg speciation in the environment described above was made possible by methodological advances in several different areas, including: i) sample collection and handling, ii) sample preparation, iii) analyte 120 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

preconcentration, iv) chromatographic separation of Hg species, and v) sensitive detection. All of these advances have been essential for obtaining relevant data. Here, however, we focus on the subject areas ii) – iv) since they are most closely related to the analysis of Hg speciation in aqueous sample preparations. Perhaps surprisingly, the bulk of the key data on Hg in environmental media was obtained using just two core analytical methods. The first method entails measuring HgII as Hg0 following a reduction process and preconcentration via gold amalgamation (23). The second method quantifies MMHg, usually by aqueous ethylation followed by purge and trap of the derivatized MeHg+ with gas chromatographic separation from other volatile Hg species (24). The ability of these methods to measure Hg in different environmental media – air, water, and solids – depends on coupling them with appropriate methods of preparing samples for analysis, preconcentrating analytes, and on having sufficiently sensitive instruments for detecting the Hg. Measurement of Hg0 is the key quantitation step in several widely used methods of analyzing total Hg, labile HgII species, and Hg0 itself. Total Hg is commonly measured by reducing to Hg0 the products of exhaustive oxidation of Hg in environmental samples (23). Such methods can be applied to different physical fractions of water or sediment samples, but cannot make chemical distinctions between the main forms of Hg present within the samples or fractions themselves. In addition, operationally-defined, labile HgII fractions have been measured by adding a reductant such as SnCl2 directly to water or sediment samples (12). Finally, sparging samples and trapping Hg0 is a direct speciation method used in studies where gaseous forms of Hg need to be quantified (7). The distillation/ethylation gas chromotagraphy (GC) method (25), which is the only widely-used method for assaying a particular ionic mercury species at ambient environmental levels, includes multiple steps (Figure 3A). As implied by the schematic of MeHg+ complexation equilibria in natural waters (Figure 2), it is necessary to first release the ion from strong organic complexes in the sample matrix before the preconcentration, chromatographic, and detection steps in analysis. Since accurate quantitation requires extremely high instrumental sensitivies to measure Hg species at the sub-parts per trillion levels found in the environment, both of the above methods usually quantify gaseous forms of Hg using atomic fluorescence spectrometry (AFS) or inductively coupled plasma mass spectrometry (ICP-MS). The main steps in using the distillation/ethylation-GC method to analyze dissolved MMHg closely parallel those in the alternative methodology described herein (Figure 3B). Both first steps, distillation and thiourea-catalyzed solid-phase extraction (TU-SPE), effectively remove the analyte(s) from the sample matrix. TU-SPE also affords a substantial factor of preconcentration that is deferred to the second step in the standard method. Both second steps concentrate and trap the analyte(s) for injection into the chromatographic system. Both third steps, gas or ion chromatography (IC), separate the analyte(s) from other Hg species. Finally, both detection steps quantify Hg in a cold vapor stream using the same detectors. The mercury-thiourea complex ion chromatography (HgTU-IC) system described below is used during the second through fourth steps in Figure 3. 121 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Figure 3. Main steps in dissolved MMHg analysis by A) distillation/ethylation-GC and B) thiourea-catalyzed solid-phase extraction/Hg thiourea complex-ion chromatography.

The differences between the chemistries of distillation/ethylation-GC and TU-SPE/HgTU-IC are evident from the ligand substitution reactions of aqueous Hg2+ amd MeHg+ complexes, both present in the natural sample and formed during the process of analysis. By manipulating the ligands coordinating each Hg species, the Hg2+ and MeHg+ can be i) removed from the original sample matrix, ii) pre-concentrated, and iii) separated or derivitized prior to detection. The reactions occurring in the sample matrix can be conceptualized as the competition for Hg between a strong natural ligand, such as a thiol moiety in dissolved organic matter (HSRDOM), and the ligands (AL) introduced in order to aid in analysis. The reactions for Hg2+ and MeHg+ with a monodentate analytical ligand can be written as:

Depending on the relative Bronsted basicities of AL and , the favored equilibrium state of the Hg species in these reactions may have significant pH dependence, making the manipulation and control of this variable central to the analytical process. 122 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

MMHg Analysis by Distillation/Ethylation-GC

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

The standard method of measuring MMHg in natural water samples, as described in United States Environmental Protection Agency (USEPA) Method 1630 (25), involves an elaborate sample preparation before the GC-based speciation and Hg detection steps. The most common sample preparation entails distillation of an acidified sample, which releases MeHg+ from the strong Hg-binding sites in dissolved organic matter (DOM) that can cause the sample matrix to interfere with direct ethylation. In DOM-rich samples, recoveries of MeHg+ internal standards can be 80% or lower with direct ethylation (26). The key reactions in water vapor distillation are ligand exchange with the analytical ligand, typically a monovalent anion such as chloride or bromide:

followed by bubbling with N2 to effect phase transfer:

and then trapping by dissociation in the condensed distillate:

It should also be noted that under the conditions in the distillation vessel, the predominant halide complexes of Hg2+ – [Hg(AL)3]− and [Hg(AL)4]2− – are themselves anionic. Thus, the distillation acts as a quite effective, if incomplete, isolation step for MeHg+. After buffering the distillate near pH 4.5, sodium tetraethyl borate is added to generate the volatile methyl ethyl mercury species (MeHgEt) from MeHg+, while any Hg2+ in the distillate reacts to form diethyl mercury (Et2Hg):

The resulting volatile alkylmercurials are purged from the ethylation vessel with N2 and preconcentrated onto a Carbotrap. Later, they are desorbed into an Ar gas stream by heating the trap and the MeHgEt separated from Et2Hg by GC. Post separation, the volatile Hg species are detected either by ICP-MS or by AFS, after pyrolysis to Hg0. Distillation/ethylation-GC is the standard method for determination of MMHg in environmental work due to its high sensitivity and low method detection limit of 0.01-0.02 ng/L for water samples (25). Note that while the distillation step helps to overcome interference by DOM, it is tedious, time-consuming, and limits the sample size that can be easily processed to ~45-60 mL (25). This keeps the detection limit higher than is desirable for marine and large lake ecosystems, although a detection limit of 0.003 ng L–1 has been achieved with speciated isotope dilution (27). 123 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Recognizing this challenge, numerous researchers have looked and are searching for an alternative way to isolate and preconcentrate dissolved MeHg+. A major focus of work on Hg speciation analysis in this laboratory since 2005 has been to develop and refine a new method of preconcentrating MeHg+ from natural waters so that large volumes can be processed and detection limits further decreased.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Speciation Analysis by Thiourea-Catalyzed Solid-Phase Extraction/HgThiourea Complex-Ion Chromatography In contrast to the GC-based analytical processes, which employ irreversible alkylation reactions to form distinct, volatile Hg species prior to preconcentration and chromatographic separation, in HgTU-IC both Hg2+ and MeHg+ are reduced to volatile Hg0 after preconcentration and chromatographic separation. Central to this process is the reversible manipulation of the ligands coordinating the Hg species. In these reactions, the key analytical ligand is TU, which by virtue of its own neutral charge, forms complexes with Hg2+ and MeHg+ (Figure 4) whose respective charges reflect the central metal ion and thus can be separated on the basis of this difference (28).

Figure 4. Structures of Hg(TU)22+ and MeHgTU+complexes.

Thiourea-Catalyzed Solid-Phase Extraction The TU-SPE process entails a series of ligand exchange reactions of Hg2+ and MeHg+ complexes with TU, the key analytical ligand. For example, to free either Hg species from the strong Hg-binding sites in DOM, water samples are acidified and amended with TU, to create conditions that favor dissociation:

Pre-concentration is then affected in an off-line preparatory step by adsorption onto thiol-functionalized resins (HSRResin) at pH ~ 4.0:

124 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

In these reactions, TU is not consumed, but its presence facilitates the quantitative adsorption of Hg species onto the resin, presumably because the ligand exchange reactions of TU complexes are more rapid. The sorbed Hg species are then eluted from the resin using strongly acidic (pH ~ 0) TU solutions (see system eluent below). This preconcentration process was therefore named “thiourea-catalyzed solid phase extraction (29).”

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

pH-Modulated Thiol-Thiourea Switch The pH dependence of the reactions comes about because the pKa of TU is less than zero, while that of a thiol is typically about 9, making it a simple matter to manipulate the relative affinities of the Hg species for the resin and the aqueous phase by altering the pH of the solution. This pH-modulated switch (Figure 5) was first employed by Shade (28) in order to permit loading of Hg from sample preparations onto the on-line thiol trap of the HgTU-IC system.

Figure 5. Thiol-thiourea switch: A) OFF (pH < 2): Thiols fully protonated and MeHg+ is complexed by TU in solution. B) ON (pH > 3) Partially-deprotonated thiols on resin outcompete TU for MeHg+.

The ligand exchange reactions of the on-line concentrator are exactly analogous to reactions (10) and (11) of the off-line TU-SPE process, with analyte loaded at pH 4 and then eluted into the system at the pH of the eluent, i.e., the 0 to 1 range. 125 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Ion Chromatographic Separation

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

In HgTU-IC, transport of complexes is retarded more than the MeHgTU+ complex due its stronger tendency to adsorb to the sulfonate moieties . Of course, when contacting an aqueous solution, these sites on the resin on the resin bed are matched by counter ions. In the HgTU-IC system, H+ is the only cation added in more than incidental amounts. Thus, the adsorption of Hg species can be written as one of the following ion exchange reactions with H+:

As a result, the retention of Hg complexes on the column must be pH-dependent. By optimizing the pH of the eluent, complete (baseline) separation between the leading MeHg+ and trailing Hg2+ peaks can be achieved. Post-Column Chemistry and Hg Detection To obtain high-sensitivity detection, the separated Hg species are transferred from the mobile phase to an Ar carrier gas stream using a multi-step, on-line reaction process. First, the difficult-to-reduce MeHg+ is oxidized to Hg2+ by a post-column reaction:

The same oxidation step also breaks down the TU, which otherwise would inhibit Hg2+ reduction. After quenching excess oxidant and neutralizing the acid, the Hg2+ is reduced to Hg0 under alkaline conditions before the sample stream is fed into a gas-liquid separator (GLS), where Hg0 is stripped into the gas phase and the cold Hg vapor carried to the detector:

The cold vapor generated by the HgTU-IC system is compatible with detection by atomic fluorescence after drying the carrier gas stream (28) or directly by ICPMS as shown here (Figure 6).

Key Changes in the HgTU-IC System The fundamental chemistry and procedures for using the HgTU-IC methodology have been published in papers by this group (28, 29) and by Shade (30). The methodology has been adopted in toto in a few instances (31) and other groups (32) have adopted major aspects of the HgTU-IC chemistry. Several analysts have made use of the pH-modulated thiol-thiourea switch to elute MeHg+ from thiol resins embedded in diffusive gradients thin film (DGT) gels (33–35). 126 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

In this lab, problems with the original post-column chemistry arose in 2009 and were not fully resolved until recently. The remainder of this chapter describes work that we have done to transform the system into a reliable and accurate tool for Hg speciation analysis and deepen our understanding of its operation. The key changes from the original system can be summarized as: • •

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

• • •

Substitution of a commercially-available thiol resin for the customsynthesized resin used originally, Adjustment of [H+] in the mobile phase in order to better separate species of low charge, including MeHg+ and EtHg+, Substitution of KBrO3 for H2O2/UV irradiation in the oxidation step (30), Substitution of NaBH4 for alkaline Sn2+ in order to avoid formation of particles that caused perturbations in the baseline, Coupling to ICP-MS detection to permit speciated isotope dilution.

A complete description of the system configuration, its operating conditions, and essential background chemistry are presented below.

System Components and Reagents Upstream of the detector, the current system components and configuration (Figure 6) are similar but not identical to those employed in earlier work in this lab (28, 29) and at Quicksilver Scientific (30). The physical parts of the system include (1) an isocratic high pressure liquid chromatography (HPLC) pump (Chromtech Series III), (2) a 2-position, 10-port sample injection valve (Rheodyne), (3) a 4×50-mm ion chromatography guard column (Dionex CG-5A), (4) an oxidation loop, (5) an oxidant-quenching loop, (6) an acid-neutralization loop, (7) a Hg reduction loop, (8) a custom borosilicate glass, gas-liquid separator (GLS) (Allen Glass, Boulder, CO), and (9) a Hg detector (Agilent 7500S ICP-MS). Peristaltic pumps are used for loading a) samples (LP) and b) reagents (RP) and for c) draining the waste solution from the GLS (WP) (Figure 6). The custom medium-pressure, concentrator/thiol trap (TT) and stock sample injection loop (SL) are both located on the injection valve. Because the acid content of the eluent ([HClEL]) was relatively high, all mechanical components of the system have an all-PEEK flow path and all tubing is made of PEEK (high pressure end of the system) or Teflon PFA (low-pressure end). Aqueous solutions containing analytes are injected into the high-pressure end of the system either via the SL or TT mounted on the injection valve (Point 2 in Figure 6). Separation of the species occurs on the ion chromatography column (ICC; Point 3 in Figure 6), which contains a mixed resin designed for analyzing trace metals. The first post-column step in the on-line reaction system is oxidation of TU by bromine monochloride (BrCl) in the oxidation loop (Point 4 in Figure 6). ) with the HCl in the The BrCl is formed upon mixing of the KBrO3 solution ( post-column acid (HCl) stream and the mobile phase. The BrCl also demethylates 127 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

MeHg+ to Hg2+. In the antioxidant loop (Point 5 in Figure 6), excess oxidants are quenched by the sodium ascorbate solution (Asc), while any hydrophobic oxidation byproducts, mostly S80, are kept in solution by the Triton XTM in the reagent. Next (Point 6 in Figure 6), the acids in the sample stream are consumed by introducing a strong base (KOH) with the resultant heat being removed as the neutralization loop passes through an ice bath. The final reaction step (Point 7 in Figure 6) is the reduction of Hg2+ to Hg0 after mixing with the alkaline borohydride ). The mobile phase, now much altered from the original eluent, solution ( passes through a gas liquid separator (Point 8 in Figure 6), where the Hg0 is stripped into the argon stream and carried to the detector (Point 9 in Figure 6).

Figure 6. Configuration of the current generation HgTU-IC system. See Table 2 for reagent compositions. (see color insert)

The particular set of post-column reactants used here (Table 2) have been extensively optimized and found to yield stable baselines and consistent sensitivities day to day. As will be discussed in more detail below, different eluents should be used in measuring MMHg and HgII, with the eluent for the latter analyte remaining unchanged from the original recipe.

Reagent Cleanup In order to maintain low blanks and chromatographic baselines, it is essential to employ reagents that are as free of Hg as is practicable. In general, we find that reagent grade chemicals are adequate. As the system reagents make substantial use of HCl, we generally purchase the trace metal grade of this acid (Fisher). However, when batches of HCl have problematical Hg blanks, we clean up 2 M solutions by adding NaBH4(s) and sparging with Ar for 1 hour or more. 128 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Table 2. Optimized System Reagents (by Analyte)a HgII

MMHg

Temperature

20

20

Flow

0.5

0.5

[HCl]

1.0

0.1(0.05)b

[HAc]

1.75

1.55(1.6)b

[TU]

0.15

0.15

Flow

0

0.25

[HCl]

--

2.5

Temperature

40

40

Flow

1.00

1.00

[KBrO3]

0.17

0.17

Temperature

20

20

Flow

0.25

0.25

[NaAsc]

1.0

1.0

[Triton-X]

1%

1%

Temperature

0

0

Flow

0.25

0.25

[KOH]

4.5

4.0

Temperature

20

20

Flow

1.00

1.00

[KOH]

1.0

1.0

Eluent

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Post-column acid (HCl)

Oxidant (

)

Antioxidant (Asc)

Base (KOH)

Reductant (

)

[NaBH4]

0.005 min–1,

All values of T in °C, Flow in mL and [X] in M units. used for greater MeHg+/EtHg+ separation. a

0.005 b

Low [HCl]EL formula

Successful operation of the system absolutely requires the use of TU that is unoxidized and has been cleaned of contaminating Hg. Upon purchase, even reagent grade TU must be examined to be sure it is free of breakdown products, which give off a strong sulfur smell or have a yellow color, and stored in a freezer. 129 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

In our experience, reagent grade TU always contains excessive Hg. Thus, we clean stock solutions (1 M TU) by adding 30-g pre-cleaned DOWEX 50W-X8 (100-200 mesh) per liter of TU solution and stirring for 1 h. Cleaned TU stocks can be stored at 4 °C for at least 14 days, or kept frozen, although we normally prepare eluent fresh each day. Oxidation of concentrated TU solutions can be detected by testing for the formation of white, S-containing polymers/particles after buffering the solution to pH 5. As Cu2+ catalytically oxidizes TU, the DOWEX must be cleaned of metals. First, DOWEX is washed by stirring in 1 M HCl for 1 h. After filtering the resin and rinsing with high-purity deionized water (DIW), it is then resuspended in 1 mM EDTA solution and neutralized with KOH. After 1 h of further stirring, the DOWEX is filtered and rinsed while on the filter with DIW. Finally, the clean TU stock is stored at 4 °C until use Additionally, we note that the thiol resin can also contain ppb levels of Hg. Most Hg can be removed by shaking batches of resin overnight in a solution of H2SO4 (2 M) and TU (150 mM) and then rinsing with DIW. Prior to use, we pump 20 mL of cleaning solution (2 M HCl + 150 mM TU) through the concentrator.

Preconcentration Using Thiol-Functionalized Resins The ability to efficiently concentrate Hg2+ and MeHg+ on thiolated resins using the pH-modulated, thiol-TU switch is an integral part of the original system that we still exploit both in off-line preconcentration from water samples and in trapping analytes in on-line concentrator prior to injection. In our earlier HgTU-IC work, we always employed a custom-synthesized, thiol-functionalized, divinyl benzene resin (28, 29). Recently, we have found that SiliaMetS®-Thiol (Silicycle, Montreal) makes an excellent replacement. From a consideration of the ligand exchange reactions (10) and (11) involved in preconcentration, one expects the trapping efficiency of the resins to depend on the pH and concentration of Hg-binding ligands in the solutions being loaded. To identify the proper pH range for efficiently trapping the analytes of interest from solutions containing our primary analytical ligand, we added MeHgCl and HgCl2 to solutions buffered to pH’s ranging from 0 to 5 with Na2SO4/H2SO4 (ionic strength 0.1 to 1) and amended with TU at 150 mM. The retention of analytes by our on-line concentrator containing ~100 mg of resin varied monotonically with pH (Figure 7). Both Hg2+ and MeHg+ were trapped completely at pH above 4 and less than 15% below pH 1, with Hg2+ being trapped to a greater extent between pH 1 and 4. The postulated competition between dissolved ligands and resin-bound thiols for the analytes during sample loading – equations (10) and (11) – also suggests that the trapping efficiency of the thiol resin depends on the concentrations of Hg-binding ligands in the solutions being loaded, i.e., bisulfide, thiosalicylic acid, glutathione, etc. To test the effect of these ligands on Hg trapping efficiency, isotopically-labeled MeHgCl and HgCl2 were added to pH 8 solutions containing 1 mM of each ligand. The experimental solutions were then pumped through the 130 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

on-line thiol concentrator. Any Hg retained by the trap was then eluted directly into the on-line reaction system and quantified. Consistent with a high affinity for Hg of the thiol resin, both analytes were completely trapped from 1 mM solutions of glutathione and thiosalicylic acid. MeHg+ was completely trapped from a solution containing 1 mM bisulfide, but only 70% of Hg2+ was retained. Since most samples are acidified for preservation, normally any sulfide present in a natural sample would be volatilized prior to preconcentration. However, to ensure that none remains behind, we recommend that samples in which the presence of sulfide is suspected should be treated by acidification and bubbling to strip H2S before attempting to trap Hg.

Figure 7. Effect of sample pH on retention of Hg species by on-line concentrator (thiol trap). [TU] = 0.15 M in loading solution.

Injection of the Analytes Analytes are injected into the high pressure end of the system via either i) a sample loop or ii) a thiol trap (concentrator) mounted on the injection valve. Each method of injection has distinct advantages and limitations that make them suitable for samples from different environmental matrices. Between the two methods, a very wide range of aqueous sample preparations can be analyzed. Sample Loop To inject via the sample loop, the sample matrix must match the eluent. This is not an inconvenience, at least for certain common sample preparation types in which system eluent can be used directly (see sample preparation). Due to limitations on the volume of sample that can be added via the loop – generally 131 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

between 10 and 100 µL but we have tested up to 1 mL – this means of injection is ideal for analyzing digests of sediments and biological tissues, but TU-SPE eluates of ambient water samples require further preconcentration.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

On-Line Concentrator/Thiol Trap (TT) At present, we manually load samples onto the on-line concentrator, or thiol trap (TT), via a loading pump using a procedure adapted from Vermillion (29). It is designed to ensure i) complete trapping of analyte, ii) prevent carryover from previous injections, and iii) minimize the MeHg+ and Hg2+ blanks. For the current system (see TT in Figure 6), the concentrator comprises SiliaMetS resin held in a custom column assembled from a 30-mm piece of 1/8”-ID PEEK tubing to which medium pressure, SuperflangelessTM fittings have been attached to each end and screwed into union junctions (all connections are ¼”-28). A PEEK frit is inserted between the tube and fitting at the downstream end (www.idex-hs.com). The new column is less prone to breaking and easier to pack than the column used in the original system. Prepared samples and the three wash/rinse solutions used in the loading process – eluent ([HCl]EL = 1 M), citrate buffer (0.1 M Na3Citrate) and DIW – are all kept in clean vials in a workspace located under a HEPA filter-fan unit. All solutions are pumped from their respective vials to the TT via a FEP sipper tube (~10-cm), a section of flexible pump tubing (~20-cm of Tygon E-3603), and a piece of 0.5-mm ID PEEK tubing (~15-cm) connected to the same side of the injection valve as the TT. The total void volume of the tubing before the trap is about 0.5 mL. The injection valve is switched at appropriate times between the positions for 1) loading the trap and 2) injecting the trapped analytes. The loading sequence begins with the valve in position 2). The loading line is flushed with enough DIW to displace the eluent that the tube is kept filled with between samples. Next, a solution of citrate buffer (0.1 M) is pumped through the tubing. Once the citrate is seen to pass through the injection valve, the valve is switched to position 1) so that the buffer flows through the TT. One mL of the citrate is then pumped through the trap and immediately followed by the sample, which is buffered to pH 4 just before loading. Then, 1 mL of DIW is added to the sample vial as a rinse to push any sample remaining in the tubing onto the trap. Finally, just enough eluent to fill the loading lines is pumped in so that they are cleaned between samples. Before this eluent flush reaches the TT, the valve is switched to position 2) so the mobile phase flows through the TT and the analytes are injected. Note that the system operator must ensure that the final flush of eluent used to clean the loading line does not reach the TT before switching the valve, or it will wash the analyte into the waste. Summary Because of the flexibility afforded by the two different methods of sample injection, the system is compatible with the main sample preparations used in the analysis of Hg species in water, sediments, and biota. Samples prepared 132 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

using extraction into organic solvents, such as toluene, can be analyzed after back-extraction into pure system eluent with [HCl]EL = 1 M (see Sediment MMHg below).

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Ion Chromatographic Separation Operation of the HgTU-IC system using the original eluent recipe, which contains [HCl]EL at 1 M (28–30), was optimized to achieve baseline separation between MeHg+ and Hg2+ at a reasonably short retention time for the latter ion. This permitted both forms of Hg to be quantified in a single chromatogram. However, when operating the system in this way, it was implicitly assumed that it was not necessary to separate MeHg+ from other Hg species of low charge, since few such species were expected to be present in typical water samples from aquatic ecosystems (Table 3). For example, Hg0 is known to be present but is rapidly oxidized in the presence of TU (data not shown) and thus would be detected as HgII. EtHg+ is another form of Hg that the system as originally operated does not distinguish from MeHg+ (29). As EtHg+ has only been reported in aquatic sediments from a few locations (36), its presence was deemed to be an unlikely source of artifactual MeHg+, at least in surface waters.

Table 3. Detection of Environmentally-Relevant Forms of Hg by HgTU-IC Species

Occurrence

Fate in HgTU-IC System

Hg0

Widespread

Oxidized to Hg2+ before injection

Me2Hg

Only in seawater

Decomposes to MeHg+ in acidic solutions (37) (Untested)

MeHg+

Widespread

Well-defined peak

EtHg+

Rare

Well-defined peak near MeHg+

Hg22+

Not reported

Untested

Hg2+

Assumed equal to InHg

Well-defined peak

HgNP

Limited reports

Untested

[HCl]EL-Dependence of Retention Times To investigate how the separation of MeHg+ from EtHg+ and Hg2+ depends on eluent pH, the times of transit from injection to the start of peaks for each species were measured over a range of [HCl]EL from 0.05 to 2.0 M using our normal 50mm CG-5A guard column. The retention time (RT) of a Hg species on the column itself is the difference between the total transit time and the time of transit without 133 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

a column, which is about 3 min. The variation in transit times of the MeHg+ peak as a function of the strong acid content of the eluent was fitted by an equation of the form:

Since the RT of MeHg+ is inversely proportional to [HCl]EL over the range 2 to 0.05 M (Figure 8), it is a simple matter to adjust eluent composition so that the system operates at any RT between 1 min for MeHg+ results in an RT for Hg2+ that is too long to be a practical method of simultaneously quantifying both species. Note also that Hg2+ retention is also more sensitive to [HAc]EL (Figure 8).

Figure 8. Effects of eluent acidity on the transit times of MeHg+, EtHg+ and Hg2+ using a single 50-mm Dionex CG-5A. (see color insert) 134 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

EtHg+ is another form of monovalent alkyl mercury that can be present in natural samples, although it is rarely reported. Since HgTU-IC relies on cation exchange to separate Hg species, one does not expect much separation of MeHg+ from EtHg+. With 0.5-1 M [HCl]EL and our standard 50-mm column, the separation between is negligible, but as [HCl]EL decreases, the relative retention of EtHg+ increases until at [HCl]EL~0.05 M there is complete separation of EtHg+ from MeHg+ (Figures 9 and 10). Note that this separation chemistry differs from the reverse-phase, ion-pairing separation of MeHg+ and EtHg+ in thiourea-based eluents reported elsewhere (38).

Figure 9. Effects of column length and eluent pH on transit time of MeHg+ (left axis) and separation of MeHg+ and EtHg+ peaks (right axis). (see color insert)

Although chromatographic separation occurs mainly on the ICC, MeHg+ and Hg2+ do elute differently from the thiol trap as well. The difference in elution profiles was directly observed by analyzing samples without the ICC in place. Both Hg species rapidly elute off the thiol trap at [HCl]EL of 1.0 M, but at 0.1 M MeHg+ elutes 1-3 min before Hg2+ with the latter exhibiting a pronounced tail. These differences likely contribute to the observed peak broadening of Hg2+ relative to MeHg+ when analyzing with the ICC. 135 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Figure 10. Smoothed HgTU-IC-ICP-MS chromatograms for a sample containing 198-enriched MeHg+ and ambient EtHg+A) before B) after mathematical source deconvolution; [HCl]EL = 0.1 M; two Dionex CG-5A 50-mm columns. See Hg Detection by ICP-MS for explanation of deconvolution. (see color insert)

Column Length We also investigated the transit times of MeHg+ and EtHg+ using three different column configurations at three different eluent pH values (Figure 9). The 50- and 100-mm column data correspond to results with one CG-5A guard columns or two in series, while the 250-mm data corresponds to results with an analytical column containing the same stationary phase (Dionex CS-5A). As expected, transit times increase with column length, with the trend being most linear at the highest pH tested (Figure 9). The separation of EtHg+ and MeHg+ also increases with column length, but the trend is less than linear above 100-mm. Getting good separation of MeHg+ and EtHg+ requires using 100-mm of column at [HCl]EL = 0.1 M (Figure 10). Figure 10A shows the individual isotopic signals measured (i.e., 198Hg-202Hg). Figure 10B displays the post-deconvolution data where “Hg198” represents signal from the [Me198HgCl] source and “HgAMB” represents the signal from the ambient EtHg+ source. The decovoluted chromatogram (Figure 10B) shows a clear separation of MeHg+ and EtHg+. See Hg Detection by ICP-MS for a full explanation of deconvolution. 136 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Fluctuations in [HCl]EL Chromatograms typically have a non-zero baseline due to a combination of instrument noise and the unavoidable presence of trace Hg2+ in the eluent and postcolumn reagents. We have observed that not only does [HCl]EL control retention times, but since the retention of Hg on the column is a cation exchange reaction – equations (12) and (13) – and H+ is the main mobile phase cation, fluctuations in [HCl]EL can perturb the baseline of chromatograms. When injecting via the sample loop, noticeable baseline perturbations can be avoided by carefully controlling the composition of the sample injected. For example, blanks and standards injected via sample loop in matrices that match the mobile phase show no “solvent dips” and peaks are highly symmetrical. But, injection of samples with matrices not matched to the eluent inserts a slug of solution with differing [H+] into the eluent stream. When a slug with low [H+] is injected, a brief (~15 s) dip in the signal baseline occurs at zero RT. Similarly, if a slug of sample with high [H+] is introduced, at zero RT the baseline is briefly raised due to the resulting perturbation in the partitioning of Hg2+ between the mobile phase and the ICC. Our qualitative observations suggest that the magnitude of the fluctuations is proportional to the background Hg2+ in the eluent. Some baseline perturbation also results from injecting analytes via the thiol trap, since loading the trap requires that the pH of the prepared sample be buffered to pH 4, i.e., 3-4 units higher than that of the eluent. While it is possible to decrease the differences between the composition of the solution filling the trap and the eluent stream, it is not possible to reliably eliminate the dip without going to great expense to reduce Hg2+ background levels in reagents.

Effects of Hg-Binding Ligands Just as injection can introduce a fluctuation in [HCl]EL, it is possible to introduce into the eluent stream a sample slug containing Hg-binding ligands that could also perturb the interaction between the mobile phase and Hg2+ sorbed to the ICC. However, as dispersion mixes the sample slug with the eluent, any introduced ligand will become increasingly dilute and less able to bind Hg2+ in the face of competition from the high concentrations of TU and H+ in the eluent. If such a ligand did outcompete TU under those conditions, it could pull Hg2+ off of the ICC and ultimately cause a brief increase in the signal baseline that might be difficult to distinguish from a peak. Samples containing Hg-binding ligands including 1 mM cysteine, glutathione or thiosalicylic acid were injected, via sample loop, into the on-line system and through the ICC at [HCl]EL = 0.1 M. None of the ligand injections caused baseline perturbations. When injecting a sample loop containing more TU than is in the mobile phase, a baseline dip resulted due to extra oxidation demand/incomplete TU oxidation.

137 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Post-Column Chemistry

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Original HgTU-IC Post-Column Chemistry The post-column redox chemistry of the HgTU-IC system employs the same two main reaction steps – oxidation of all Hg species to Hg2+ followed by reduction to Hg0 − as other flow-injection methods for total Hg analysis (39), but its reactant fluxes and compositions are adjusted to also fully oxidize the TU in the mobile phase and to attain each process’s requisite pH. Although smooth baselines and high sensitivities were routinely achieved in earlier work using H2O2/UV-oxidation and reduction by alkaline Sn2+ (29), subsequently we observed formation of two types of fine particles within the post-column reduction loop that randomly disturbed the baseline. The grey/black particles observed were likely Sn(OH)2(s), as stannous hydroxide becomes supersaturated within the reduction loop. The white precipitates were likely elemental sulfur or formamidine disulfide, which are known products of TU oxidation with limited solubilities (40). We had some success in avoiding particle formation by carefully selecting high quality reagents. The precipitation of Sn(OH)2(s) could be mitigated to some extent by i) neutralizing the acid in the mobile phase after oxidation prior to mixing with the highly alkaline Sn2+ reductant (30), and by ii) raising the hydroxide concentration to 10 M (unpublished results). In addition, adding Triton-X to the antioxidant reduced formation of S-containing particles (6). However, some fine particles were always formed after several hours of operation using these chemistries. In order to avoid this problem, alternative chemistries for the oxidation/reduction steps were investigated. In particular, oxidizing the S-II in TU more completely and finding an alternative to reduction by alkaline Sn2+ were deemed essential.

Oxidation of Thiourea That TU was not completely oxidized in the original HgTU-IC method was shown in an experiment conducted using an alternative reductant, 20% SnCl2 in 20% HCl (41). When following UV/H2O2 oxidation with reduction by acidic SnCl2, we observed no formation of Hg0 from MeHg+, implying that either MeHg+ had not been oxidized or that enough TU remained that it could inhibit the reduction of the Hg2+ formed from MeHg+ oxidation. Since reduction clearly occurs with alkaline Sn2+ and since MeHg+ is stable under alkaline conditions, while TU is not (42), this result implies that hydrolysis of incompletely-oxidized TU permitted Hg reduction to proceed in the original HgTU-IC method. Rather than attempting to further optimize the H2O2/UV chemistry, we tested oxidation by bromate. KBrO3 is commonly employed as a precursor of the BrCl oxidant used to measure total Hg in water (23), and was adopted in in HgTU-IC, Shade’s update of HgTU-IC chemistry (30). When using yellow-orange dissolved Br2 is visible after the oxidant stream mixes with the TU-containing mobile phase, as indicated in equation (17). Although the formation of white S-containing particles was diminished at the published ratio of 138 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

bromate-to-TU fluxes (0.74: 1) (30), some very fine white particles still appeared in the neutralization loop after running the system for several hours. To determine how much bromate is necessary to completely oxidize the S-II in TU, the system was operated using acidic SnCl2 as the reductant in order to avoid alkaline hydrolysis of TU. Various bromate concentrations in the oxidant were tested using constant flow rates of all reagents and constant composition of eluent (1 M [HCl]EL, 150 mM [TU]EL) and reductant (20% SnCl2 in 20% HCl). Peak areas for 100 pg MeHg+ standards were measured and are reported here (Figure 11).

Figure 11. HgTU-IC-ICP-MS system response to injected MeHg+ while varying bromate/TU molar flux ratio with reduction by acidic Sn2+.

The dependence of the system sensitivity on bromate flux observed in the experiment (Figure 11) shows that oxidation of TU’s S-II is essentially complete when the bromate and eluent streams are mixed at a 1.6: 1 molar ratio of bromate to TU fluxes. This ratio agrees closely with the reported stoichiometry of TU when the latter is present in excess (43): oxidation by

Subsequently, we operated the system using a : TU flux ratio that slightly exceeds the Simoyi stoichiometry and with the oxidation coil immersed in a 40 °C water bath to increase the reaction rate. Excess oxidant is needed to ensure /Br2/BrCl, NaAsc is injected into complete TU oxidation. To quench the extra the flow and allowed to react in the antioxidant loop located before the reduction step (Figure 6). 139 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Borohydride Reduction

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

While not used as widely as acidic SnCl2, borohydride has been employed as a post-column reductant in at least one reverse-phase-HPLC Hg speciation system (44), and suggested as a suitable replacement for alkaline Sn2+ in the HgTU-IC system (30). We found that at reduction loop transit times of ~30 s, NaBH4 is an effective reductant at concentrations as low as 5 µM and that system sensitivity increases only 20% when its concentration is raised by five orders of magnitude (Figure 12).

Figure 12. Effects of concentration and residence time within the reduction loop on HgTU-IC system sensitivity to MeHg+.

Neutralization of Acid Prior to Reduction Loop Although is an effective reductant under acidic conditions, the H2(g) bubbles formed at low pH add fluctuations to the signal. Thus, only alkaline NaBH4 is a suitable reagent in the HgTU-IC system. Since sensitivity varies inversely with pH, a balance was found to maximize signal while safeguarding against the formation of H2(g) bubbles. To do this, the acid flux in the eluent was matched by the base addition, which was then mixed with the 1 M KOH in the is followed with Hg2+ reductant. When the complete oxidation of TU by , the HgTU-IC system can operate with no particle formation reduction by over long periods (>10 h). An advantage this version of HgTU-IC has over other LC methods is the nature of the waste stream it produces, i.e., it contains no organic solvents or metals. By adding a small amount of acetone and bring the waste to pH 4, one , leaving acetone, acetate, Triton-X, Br–, Cl–, K+, can consume any excess + Na , ascorbate oxidation products, and urea in the waste. Also, the final pH of this waste is much less alkaline than waste from the previous HgTU-IC chemistries, making neutralization easier and less expensive. 140 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Summary

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

We have found the post-column chemistry described herein to be very stable, free of baseline pertubations, and at least as sensitive as the best results obtained with earlier versions of the HgTU-IC system (29). The fluxes of acids and bases are more carefully balanced (Table 4) to yield optimal sensitivity while also simplifying neutralization of wastes. The oxidant flux is also substantially higher than in previous versions in order to eliminate the formation of elemental sulfur particles.

Table 4. Acid-Base and Redox Balances in Different HgTU-IC System Chemistries Analytes

MMHg + HgII

MMHg + HgII

MMHg + HgII

System Version

2005 (28)

2007 (29)

2008 (30)

Proton Balance (Fluxes in meq

HgII

MMHg This work

min–1)

Eluent HCl

0.45

0.5

0.5

0.5

0.05

Eluent HAc

0.87

0.875

1.3

0.875

0.775

Oxidant HCl

0

0

0

0

0.625

Oxidation products

0.2

0.2

0.14

0.15

0.15

–3

–1.13

–1

Base Reductant

–5.7

–5.8

–0.75

–1

–1

Net protons

–4.18

–4.23

–1.81

–0.60

–0.4

Electron Balance (Fluxes in meq min–1) TU

0.8

0.8

0.56

0.6

0.6

Oxidant

–0.68

–0.52

–0.25

–0.85

–0.85

Net electrons

0.12

0.28

0.31

–0.25

–0.25

Ascorbate

0.28

0.28

0

0.3

0.3

Hg Detection by ICP-MS This work is the first documenting the use of HgTU-IC with cold vapor generation and ICP-MS detection; all previous work was performed using CV-AFS (28) or ICP-MS analysis of HgTU-IC eluent without post-column reaction (32). Just as shown by workers using ICP-MS with ethylation-GC, the HgTU-IC-ICP-MS system yields individual chromatograms for each of the different Hg isotopes (Figures 10A and 13A). Note that in this chromatogram, the natural abundance of the Hg isotopes in the ambient MeHgCl standard is perturbed 141 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

by the added Me198HgCl internal standard. Thus, as has been previously shown for THg and MMHg analysis (41), employing ICP-MS detection allows one to perform species specific stable isotope dilution, which enhances precision through the use of internal standards and facilitates the use of isotopically-enriched tracers to assay rates of species transformations in environmental media.

Figure 13. Smoothed HgTU-IC-ICP-MS chromatograms containing a mixture of ambient and 198Hg-enriched MeHg+ A) before B) after mathematical source deconvolution; [HCl]EL = 0.1 M. (see color insert) An important benefit of using ICP-MS for the HgTU-IC system is that it enables the direct comparison of peak shapes of analyzed Hg species in unknown samples with those of known isotopically-labelled internal standards. To make such comparisons, we perform a mathematical deconvolution using the ICP-MS counts for each isotope at every time point (0.1 s resolution) in the chromatogram (Figure 13A) prior to the integration of peak areas, a step that is not commonly done (41, 45). The deconvolution yields chromatographic traces expressed in terms of counts of ambient Hg (HgAMB) and those of the labelled internal standard (Hg198) (Figure 13B). While the trained eye can readily detect the perturbation of the natural abundance of the Hg isotopes from the relative size of the Me198Hg+ and Me202Hg+ peaks, the consistency of the peak shapes and transit times are much more evident after the deconvolution. In natural samples, we recommend using linear regression of the individual time points from the deconvoluted chromatograms to quantify the similarity of the ambient and internal standard peaks. In this case, the correlation coefficient (R2) of the data points during the peak was 0.992. If there were an ambient Hg species 142 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

that overlapped with MeHg+ under the system operating conditions, such as EtHg+, there would be a clear deviation from linearity over the latter part of the peak and a poor correlation coefficient. The regression can be used as a tool to ensure that no other Hg species is overlapping with MeHg+ under the operating conditions.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Measuring Dissolved MMHg and HgII by TU-SPE/HgTU-ICICP-MS To analyze HgII and/or MMHg at low levels in natural water samples, an off-line TU-SPE step is necessary to pre-concentrate Hg species and reduce the amounts of undesirable matrix components (DOM and other metals) injected into the on-line system. Our current off-line TU-SPE procedure is almost unchanged from the original (29), with the exception that we now use the Silia MetS-Thiol resin and add isotopically-labelled internal standards in the leaching step. Briefly, 20- to 1000-mL subsamples of previously filtered and acidified primary samples are weighed into clean borosilicate glass vials. Stock solutions containing clean TU and isotopically-labelled MeHgCl and HgCl2 internal standards are added to attain the desired concentrations, e.g., TU at 10-40 mM. Typically, the capped vials are then leached overnight at room temperature, although shorter leaching can be used (29). Samples can be buffered to pH 4 by addition of an appropriate volume of Na3Citrate stock (0.75 M) either before leaching or just before the TU-SPE step. On the day off-line TU-SPE is performed, we prepare TU-SPE columns by slurry-packing ~100-mg of thiol resin into each of several borosilicate glass chromatography columns (Kontes, 1×5-cm) and placing a small wad of glass wool on top of the resin bed. The packed column is then washed with 20 mL of cleaning solution (2 M HCl and 150 mM TU) and rinsed with 10 mL of DIW. The sample loading sequence begins with pumping: i) ethanol (10 mL), ii) eluent (10 mL at 1 M [HCl]EL), iii) DIW (10 mL), and iv) sodium citrate buffer (1 mL of 0.75 M) through the packed column. Then, the leached and buffered samples are pumped through the column. Following sample loading, the adsorbed analytes are eluted with 4 mL of eluent (1 M [HCl]EL), which is then kept frozen until analysis. Note that by preconcentrating the Hg species off-line, relatively little of the DOM and other metals from the original water sample end up in the prepared sample that is loaded on-line. Much of the DOM flows through the resin column without being adsorbed and most of the DOM that does adsorb is left on the resin when the Hg is eluted. This cleanup enables one to load samples on-line without plugging the TT frit and removes solutes that might degrade the performance of the on-line thiol resin or ICC. Performance Metrics To test method performance for dissolved MMHg and HgII, detection limit (MDL) studies were performed. The MDLs were determined by analyzing seven replicate samples cotaining 1 pg MeHg+ and 5 pg Hg2+. The standard deviation 143 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

from the replicates was multiplied by the student’s t-value appropriate for 99% confidence level to calculate the MDLs. The Hg stock solutions were diluted into 40 mL samples and TU-SPE was performed as described above. Both analytes were measured in a single chromatogram, while operating the system at 1 M [HCl]EL. We expect the variability in MMHg peak integrations to be greater under these conditions relative to operating at 0.1 M. The resultant MDLs are 0.003 ng L–1 for MMHg and 0.01 ng L–1 for HgII. Synthetic samples were created in order to test the effect of DOM on Hg preconcentration by TU-SPE. A stock solution containing ambient MeHg+, Hg2+, and Suwanee River DOM (IHSS) in the mass proportions 1 ng: 5 ng: 10 mg was added to Hg-free spring water at three different dilutions and internal standards added for both analytes. All samples in the study were run in triplicate; reported concentrations are the average values after correcting for recovery of the internal standards using standard isotope dilution calculations (45). The recoveries of the ambient MMHg and HgII from 30 mL samples, calculated as percent of the value expected based on the dilution of the sample, was unaffected by DOM between 2 and 20 mg L–1 or by the presence of EDTA at 1 mM (Table 5). The average recoveries of internal standards were 93% and 92% for MMHg and HgII respectively.

Table 5. Concentrations and Recoveries of Hg Species from Synthetic Samples Containing Suwanee River NOM and EDTA Using TU-SPE / HgTU-IC-ICP-MS [DOM]

2 mg

L–1

8 mg L–1

20 mg L–1

Units

MMHg

HgII

MMHg w/EDTA

HgII w/EDTA

ng L–1

0.203 ±0.001

1.038 ±0.043

0.194 ±0.004

0.987 ±0.005

Recovery

101.7%

103.8%

97.2%

98.7%

ng L–1

0.773 ±0.006

3.952 ±0.054

0.792 ±0.010

3.849 ±0.148

Recovery

96.7%

98.8%

99.0%

96.2%

ng L–1

1.992 ±0.034

9.954 ±0.132

2.026 ±0.035

9.636 ±0.077

Recovery

99.6%

99.5%

101.3%

96.4%

Since in this method, the recovery of Hg from the samples depends on the presence of TU, it was deemed important to check whether the method’s performance is affected by the presence of a known TU oxidizer commony found in natural waters, Cu2+ (46). At 1.5 mg/L Cu2+, there was a significant amount of TU oxidation as evident from the clouding of the samples during leaching at pH 4. The addition of Cu2+ without EDTA did not affect the recovery of MMHg, but 144 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

did reduce the HgII recovery to an average of 86.2%. In the samples containing both 1.5 mg/L Cu2+ and 1 mM EDTA there was no clouding during the leaching step and recovery of both MMHg and HgII was quantitative. Based on these results, it is recommended that samples known to or suspected of containing high levels of Cu be amended with enough EDTA to bind it.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Sediment MMHg To analyze total MMHg in aquatic sediment samples, we employ a modified version of Bloom’s digestion/extraction technique (22). In the original method, an H2SO4/CuSO4/KBr solution is mixed with sediment to leach sediment-bound MeHg+ into solution as the neutral complex [MeHgBr]0. A simultaneous extraction with dichloromethane (DCM) removes the complex from the aqueous phase. After the extraction, the DCM phase is subsampled into a new vessel containing deionized water and the DCM allowed to evaporate, leaving behind the MeHgBr0 dissolved in deionized water. The resulting aqueous sample can be analyzed by ethylation-GC. For coupling with HgTU-IC, the same procedure could be used with eluent replacing water in the last step. However, it is convenient to make an additional substitution of toluene for DCM in order to work with a less volatile solvent. The resultant sample preparation is ready to be buffered and loaded onto the HgTU-IC system. To apply this method, ~0.2-0.5 g of sediment is shaken for 1 h with a mixture of 5 mL of 18% (w/v) KBr + 5% (v/v) H2SO4 and 1 mL of 1 M CuSO4. An appropriate amount of isotopically-enriched MeHgCl internal standard is also added at this stage. The combination of acid and ligand leaches MeHg+ into solution, forming the neutral MeHgBr0 species. Then 10 mL of toluene is added to the mixture and shaken for 1 h to extract the neutral MeHgBr0 species into the toluene phase and drive the desorption reactions to completion. After mixing, the sample is centrifuged to break up any emulsion that forms and to separate suspended solids from the toluene phase. Next, 80-90% of the toluene is transferred into a new vessel containing 5 mL of eluent and shaken for 1 h. The [MeHgBr]0 dissolves back into the aqueous phase. There the Br– anion is replaced by TU to form the charged MeHgTU+ species, causing all of the [MeHgBr]0 to be transferred from the toluene into the aqueous eluent. A sub-sample of the eluent can now be injected via sample loop or buffered and loaded onto the thiol trap as described above.

Other Sample Preparations Total Dissolved Hg For THg analysis the sample oxidation step adapted from USEPA Method 1631 is used (23). The sample is brought to 1-5% BrCl in order to oxidize DOM and allowed to react for at least 48 h. To quench excess BrCl, hydroxylamine 145 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

(NH2OH) solution (30%) is typically added at 15 µL per 10-mL of sample. Finally, the digested sample is brought up to 50 mM TU to prevent the potentially substantial adsorption of Hg2+ to the sample vial and the sample introduction system that can occur once the sample is buffered to pH 4-5 for on-line loading. The loading procedure is the same as described above.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Total Sediment Hg A wide variety of methods for digesting sediments for total Hg are known. Essentially all can be adapted for use with HgTU-IC by quenching excess oxidant with NH2OH or NaAsc, adding TU to keep the Hg2+ in solution, and buffering with citrate or acetate at pH ~4 just prior to loading.

Biological Tissue Digestions A simple and effective procedure to extract MeHg+ and Hg2+ from biological tissues by digesting them overnight in eluent at 60 °C has been reported by Shade (30).

Recommended System Operating Conditions Conditions for Hg Speciation Analysis The key system variables that one can tune to optimize the tradeoff between peak separation and analysis time are i) eluent proton concentration, or [HCl]EL and ii) length of the column (ICC). Since lengthening the column also raises the pressure in the system, we normally adjust [HCl]EL so that the peaks corresponding to the Hg species we need to distinguish and/or quantify i) can be resolved from each other and ii) come long enough after the injection dip to allow the baseline to be reestablished. To measure MMHg in samples containing little or no EtHg+, it is convenient to use one 4×50-mm ICC with an eluent that contains [HCl]EL of 0.1 M or less. Although this setup affords only partial separation of MeHg+ and EtHg+, when combined with the use of isotopically-enriched interal standards and ICP-MS detection one can identify the presence of EtHg+ by comparing the internal standard and ambient Hg curves in the deconvoluted chromatograms. When using AFS detection, the identification of MeHg+ peaks influenced by EtHg+ is less certain. Thus, when analyzing MMHg by HgTU-IC-AFS or ETHg with either type of detector, one should either employ two 4×50-mm columns at [HCl]EL of 0.1 M or one 4×50-mm columns at 0.05 M [HCl]EL, to achieve complete separation. To analyze HgII in samples not subjected to oxidative preparations, we recommend using an [HCl]EL of 1.0 M (Table 2) to adequately separate Hg2+ from singly-charged Hg species without excessively long total analysis times. 146 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Table 6. Results of MMHg Analysis Using HgTU-IC-ICP-MS for Intercomparison Samples and Reference Materials Sample Mass (g)

HgTU-IC Value

Consensus Value

Intercomp 2014 “UB” (47)

50

0.110 ± 0.022

0.126 ± 0.080

Intecomp 2014 “SP” (47)

50

0.118 ± 0.007

0.103 ± 0.050

Intercomp 2015 “UJ” (48)

50

0.038 ± 0.002

300

0.045 ± 0.006

50

0.240 ± 0.015

300

0.240 ± 0.001

50

0.041 ± 0.001

300

0.054 ± 0.004

Dissolved MMHg (ng L–1)a

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Intercomp 2015 “LS” (48)

Intercomp 2015 “CC” (48)

0.043 ± 0.011

0.22 ± 0.087

0.049 ± 0.011

Sediment MMHg (ng g-dw–1)b BRI-1 (49)

0.32

0.168 ± 200.003

0.17 ± 0.06c

IAEA 158

0.30

1.40 ± 200.14

1.38 ± 0.27

CC 580

0.23

72.9 ± 204.5

75.5 ± 3.7

a Samples prepared using TU-SPE. b Samples prepared using H SO /CuSO /KBr/toluene 2 4 4 digestion+extraction. c Provisional value; Due to wide scatter in results, median is reported here instead of mean of 0.20.

Conditions for THg Analysis There are two main setups for the HgTU-IC system that one can use to analyze THg, i.e., Hg2+ in samples subjected to preparative procedures that exhaustively oxidize all Hg species. The choice between the two depends on the range of concentrations one needs to quantify. The simplest method is appropriate for preparations containing relatively high levels of Hg. With such samples, one only needs to dilute them into eluent and inject small volumes via the sample loop. The time of analysis can be minimized by operating without the column (ICC), since only Hg2+ is present. Samples of surface waters or other media containing low levels of Hg must be injected into the system via the thiol trap. However, this method of loading creates an injection dip that slightly precedes but is not separated from the Hg2+ peak. Thus, in order to separate the two and accurately integrate the peak, one must employ the ICC. By adjusting [HCl]EL in the range of 1.0-1.3 M, one can allow the baseline to be reestablished after the dip while minimizing the retention time of the Hg2+ peak. 147 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

Method Validation While it is conceivable that a new analytical method such as this could yield systematically different results than the standard methods if it recovered more or less of the analyte, that is not the case here. Results from participation in blind intercomparison studies for dissolved MMHg and MMHg in sediment reference materials indicate that our results with HgTU-IC-ICP-MS are all within the 95% confidence limits of the consensus or certified values (Table 6). The high accuracy of the system/sample preparation combination for biota has already been documented (30). Note that when analyzing water, excellent results were obtained for MMHg with both 50- and 300-mL samples. Since the recoveries of analytes were high even in the large volume samples, routine analysis of water samples with detection limits in the low pg L–1 range for MMHg should be possible.

Acknowledgments We are grateful for the financial support of this work provided by i) a Natural Resources Conservation Service Conservation Innovation Grant to the Iowa Soybean Association, ii) Electric Power Research Institute project EP-P30063/C14095, and iii) USDA National Institute of Food and Agriculture, Hatch project 875-913. R.K. was supported by a postdoctoral fellowship from The Camille and Henry Dreyfus Foundation, Inc.. The training in speciated isotope dilution analysis of MMHg afforded by H. Hintelmann and his research group during R.H.’s sabbatical at Trent University was enormously influential in this work.

References 1.

2.

3.

4.

5. 6.

Driscoll, C. T.; Mason, R. P.; Chan, H. M.; Jacob, D. J.; Pirrone, N. Mercury as a Global Pollutant–Sources, Pathways, and Effects. Environ. Sci. Technol. 2013, 47, 4967–4983. Watras, C. J.; Morrison, K. A.; Host, J. S.; Bloom, N. S. Concentration of Mercury Species in Relationship to Other Site-Specific Factors in the Surface Waters of Northern Wisconsin Lakes. Limnol. Oceanogr. 1995, 40, 556–565. Davis, A.; Bloom, N. S.; Que Hee, S. S. The Environmental Geochemistry and Bioaccessibility of Mercury in Soils and Sediments: A Review. Risk Anal. 1997, 17, 557–569. Watras, C. J.; Back, R. C.; Halvorsen, S.; Hudson, R. J. M.; Morrison, K. A.; Wente, S. P. Bioaccumulation of Mercury in Pelagic Freshwater Food Webs. Sci. Total Environ. 1998, 219, 183–208. Selin, N. E. Global Biogeochemical Cycling of Mercury: A Review. Annu. Rev. Environ. Resour. 2009, 34, 43–63. Fitzgerald, W. F.; Lamborg, C. H.; Hammerschmidt, C. R. Marine Biogeochemical Cycling of Mercury. Chem. Rev. 2007, 107, 641–662. 148 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

7. 8.

9.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

10.

11.

12.

13. 14.

15.

16.

17.

18.

19. 20.

Vandal, G. M.; Mason, R. P.; Fitzgerald, W. F. Cycling of Volatile Mercury in Temperate Lakes. Water Air Soil Pollut. 1991, 56, 791–803. Deonarine, A.; Lau, B. L. T.; Aiken, G. R.; Ryan, J. N.; Hsu-Kim, H. Effects of Humic Substances on Precipitation and Aggregation of Zinc Sulfide Nanoparticles. Environ. Sci. Technol. 2011, 45, 3217–3223. Pham, A. L.-T.; Morris, A.; Zhang, T.; Ticknor, J.; Levard, C.; Hsu-Kim, H. Precipitation of Nanoscale Mercuric Sulfides in the Presence of Natural Organic Matter: Structural Properties, Aggregation, and Biotransformation. Geochim. Cosmochim. Acta 2014, 133, 204–215. Skyllberg, U. Competition among Thiols and Inorganic Sulfides and Polysulfides for Hg and MeHg in Wetland Soils and Sediments under Suboxic Conditions: Illumination of Controversies and Implications for MeHg Net Production. J. Geophys. Res. 2008, 113, 1–14. Hintelmann, H.; Welbourn, P. M.; Evans, R. D. Measurement of Complexation of Methylmercury(II) Compounds by Freshwater Humic Substances Using Equilibrium Dialysis. Environ. Sci. Technol. 1997, 31, 489–495. Lamborg, C. H.; Tseng, C.-M.; Fitzgerald, W. F.; Balcom, P. H.; Hammerschmidt, C. R. Determination of the Mercury Complexation Characteristics of Dissolved Organic Matter in Natural Waters with “Reducible Hg” Titrations. Environ. Sci. Technol. 2003, 37, 3316–3322. Hsu, H.; Sedlak, D. L. Strong Hg(II) Complexation in Municipal Wastewater Effluent and Surface Waters. Environ. Sci. Technol. 2003, 37, 2743–2749. Black, F. J.; Bruland, K. W.; Flegal, A. R. Competing Ligand ExchangeSolid Phase Extraction Method for the Determination of the Complexation of Dissolved Inorganic Mercury (II) in Natural Waters. Anal. Chim. Acta 2007, 598, 318–333. Hintelmann, H. 11. Organomercurials. Their Formation and Pathways in the Environment. In Metal Ions in Life Sciences; Sigel, A., Sigel, H., Sigel, R. K. O., Eds.; Royal Society of Chemistry: London, 2010; Vol. 7, pp 365–401. Skyllberg, U.; Drott, A. Competition between Disordered Iron Sulfide and Natural Organic Matter Associated Thiols for Mercury(II)–An EXAFS Study. Environ. Sci. Technol. 2010, 44, 1254–1259. Jonsson, S.; Skyllberg, U.; Nilsson, M. B.; Westlund, P. O.; Shchukarev, A.; Lundberg, E.; Björn, E.; Bjo, E. Mercury Methylation Rates for Geochemically Relevant HgII Species in Sediments. Environ. Sci. Technol. 2012, 46, 11653–11659. Paquette, K. E.; Helz, G. R. Inorganic Speciation of Mercury in Sulfidic Waters: The Importance of Zero-Valent Sulfur. Environ. Sci. Technol. 1997, 31, 2148–2153. Jay, J. A.; Morel, F. M. M.; Hemond, H. F. Mercury Speciation in the Presence of Polysulfides. Environ. Sci. Technol. 2000, 34, 2196–2200. Bell, A. M. T.; Charnock, J. M.; Helz, G. R.; Lennie, A. R.; Livens, F. R.; Mosselmans, J. F. W.; Pattrick, R. A. D.; Vaughan, D. J. Evidence for Dissolved Polymeric Mercury(II)-Sulfur Complexes? Chem. Geol. 2007, 243, 122–127. 149 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

21. Qian, J.; Skyllberg, U.; Frech, W.; Bleam, W. F.; Bloom, P. R.; Petit, P. E. P. Bonding of Methyl Mercury to Reduced Sulfur Groups in Soil and Stream Organic Matter as Determined by X-Ray Absorption Spectroscopy and Binding Affinity Studies. Geochim. Cosmochim. Acta 2002, 66, 3873–3885. 22. Bloom, N. S.; Colman, J. A.; Barber, L. Artifact Formation of Methyl Mercury during Aqueous Distillation and Alternative Techniques for the Extraction of Methyl Mercury from Environmental Samples. Fresenius’ J. Anal. Chem. 1997, 358, 371–377. 23. Method 1631, Revision E: Mercury in Water by Oxidation, Purge and Trap, and Cold Vapor Atomic Fluorescence Spectrometry; U.S. Environmental Protection Agency, Office of Water: Washington, DC, 2002. 24. Bloom, N. S. Determination of Picogram Levels of Methylmercury by Aqueous Phase Ethylation, Followed by Cryogenic Gas Chromatography with Cold Vapour Atomic Fluorescence Detection. Can. J. Fish. Aquat. Sci. 1989, 46, 1131–1139. 25. Method 1630: Methyl Mercury in Water by Distillation, Aqueous Ethylation, Purge and Trap , and CVAFS (Draft); U.S. Environmental Protection Agency, Office of Water, Washington, D.C, 2001. 26. Hintelmann, H. Personal communication, 2007. 27. Jackson, B.; Taylor, V.; Baker, R. A.; Miller, E. Low-Level Mercury Speciation in Freshwaters by Isotope Dilution GC-ICP-MS. Environ. Sci. Technol. 2009, 43, 2463–2469. 28. Shade, C. W.; Hudson, R. J. M. Determination of MeHg in Environmental Sample Matrices Using Hg-Thiourea Complex Ion Chromatography with on-Line Cold Vapor Generation and Atomic Fluorescence Spectrometric Detection. Environ. Sci. Technol. 2005, 39, 4974–4982. 29. Vermillion, B. R.; Hudson, R. J. M. Thiourea Catalysis of MeHg Ligand Exchange between Natural Dissolved Organic Matter and a Thiol-Functionalized Resin: A Novel Method of Matrix Removal and MeHg Preconcentration for Ultratrace Hg Speciation Analysis in Freshwaters. Anal. Bioanal. Chem. 2007, 388, 341–352. 30. Shade, C. W. Automated Simultaneous Analysis of Monomethyl and Mercuric Hg in Biotic Samples by Hg-Thiourea Complex Liquid Chromatography Following Acidic Thiourea Leaching. Environ. Sci. Technol. 2008, 42, 6604–6640. 31. Oiffer, L.; Siciliano, S. D. Methyl Mercury Production and Loss in Arctic Soil. Sci. Total Environ. 2009, 407, 1691–1700. 32. Hong, Y. S.; Rifkin, E.; Bouwer, E. J. Combination of Diffusive Gradient in a Thin Film Probe and IC-ICP-MS for the Simultaneous Determination of CH3Hg+ and Hg2+ in Oxic Water. Environ. Sci. Technol. 2011, 45, 6429–6436. 33. Clarisse, O.; Hintelmann, H. Measurements of Dissolved Methylmercury in Natural Waters Using Diffusive Gradients in Thin Film (DGT). J. Environ. Monit. 2006, 8, 1242–1247. 34. Gao, Y.; De Canck, E.; Leermakers, M.; Baeyens, W.; Van Der Voort, P. Synthesized Mercaptopropyl Nanoporous Resins in DGT Probes for Determining Dissolved Mercury Concentrations. Talanta 2011, 87, 262–267. 150 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch006

35. Fernández-Gómez, C.; Bayona, J. M.; Díez, S. Comparison of Different Types of Diffusive Gradient in Thin Film Samplers for Measurement of Dissolved Methylmercury in Freshwaters. Talanta 2014, 129, 486–490. 36. Cai, Y.; Jaffé, R.; Jones, R. Ethylmercury in the Soils and Sediments of the Florida Everglades. Environ. Sci. Technol. 1997, 31, 302–305. 37. Conaway, C. H.; Black, F. J.; Gault-Ringold, M.; Pennington, J. T.; Chavez, F. P.; Flegal, A R. Dimethylmercury in Coastal Upwelling Waters, Monterey Bay, California. Environ. Sci. Technol. 2009, 43, 1305–1309. 38. Dórea, J. G.; Wimer, W.; Marques, R. C.; Shade, C. W. Automated Speciation of Mercury in the Hair of Breastfed Infants Exposed to Ethylmercury from Thimerosal-Containing Vaccines. Biol. Trace Elem. Res. 2011, 140, 262–271. 39. Leopold, K.; Foulkes, M.; Worsfold, P. Methods for the Determination and Speciation of Mercury in Natural Waters-A Review. Anal. Chim. Acta 2010, 663, 127–138. 40. Gherrou, A.; Kerdjoudj, H.; Molinari, R.; Drioli, E. Modelization of the Transport of Silver and Copper in Acidic Thiourea Medium through a Supported Liquid Membrane. Desalination 2001, 139, 317–325. 41. Hintelmann, H.; Ogrinc, N. Determination of Stable Mercury Isotopes by ICP/MS and Their Application in Environmental Studies. In Biogeochemistry of Environmentally Important Trace Elements; Cai, Y., Braids, C. O., Eds.; ACS Symp Series Volume 835; American Chemical Society Publishing: Washington, DC, 2003; Vol. 835, pp 321–338. 42. Norr, M. K. The Lead Salt-Thiourea Reaction. J. Phys. Chem. 1954, 65, 1278–1279. 43. Simoyi, R. H.; Epstein, I. R.; Kustin, K. Kinetics and Mechanism of the Oxidation of Thiourea by Bromate in Acidic Soluton. J. Phys. Chem. 1994, 98, 551–557. 44. Ilgen, R. F. G. Coupling of the RP C18 Preconcentration HPLC-UV-PCOSystem with Atomic Fluorescence Detection for the Determination of Methylmercury in Sediment and Biological Tissue. Fresenius’ J. Anal. Chem. 1997, 358, 407–410. 45. Meija, J.; Yang, L.; Caruso, J. A.; Mester, Z. Calculations of Double Spike Isotope Dilution Results Revisited. J. Anal. At. Spectrom. 2006, 21, 1294. 46. Zatko, D. A.; Kratochvil, B. Copper(II) Oxidation of Thioureas in Acetonitrile. Anal. Chem. 1968, 40, 2120–2123. 47. Creswell, J.; Metz, J.; Carter, A.; Davies, C. 2014 Brooks Rand Labs Interlaboratory Comparison Study for Total Mercury and Methylmercury ( Intercomp 2014 ); Seattle, 2014. 48. Creswell, J.; Kilner, P. I.; Davies, C. 2015 Brooks Rand Labs Interlaboratory Comparison Study for Total Mercury and Methylmercury ( Intercomp 2015 ); Seattle, 2015. 49. Creswell, J. BRI-1 Wisconsin Freshwater Sediment Reference Sample for Total Mercury and Methylmercury, Provisional Report; Seattle, 2015.

151 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Chapter 7

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch007

Fate of Chlorate and Perchlorate in High-Strength and Diluted Hypochlorite Solutions Anna Breytus,* Srinivas Prabakar, and Andrew P. Kruzic Department of Civil Engineering, University of Texas at Arlington, Arlington, Texas 76019 *E-mail: [email protected].

Hypochlorite solutions have a potential of introducing disinfection by-products, such as the oxyhalides chlorate, perchlorate and bromate into the drinking water when used for drinking water disinfection. Measurement of these by-products in various strength hypochlorite solutions has been an issue of importance in the last decade, especially in view of the current aim of the Homeland Security Department to reduce the usage of chlorine gas. Previous work identified presence of oxyhalides in both low- and high-strength hypochlorite solutions. Perchlorate is an endocrine disruptor that inhibits iodide intake by the thyroid, thus reducing the production of essential thyroid hormones. Chlorate also has several adverse effects on the blood and thyroid systems. Perchlorate is regulated in the states of Massachusetts and California. Chlorate has a health reference level established by the EPA. In addition, it has a notification and a proposed action level in the state of California. Both contaminants are also being considered for federal regulation. One of the goals of the study was to investigate the degree to which hypochlorite solutions degrade and examine the increase in chlorate and perchlorate concentrations in storage tanks. This was accomplished by comparing chlorine, chlorate and perchlorate levels in storage tanks to newly delivered solutions. Two facilities that had different suppliers were sampled. From a comparison between stored and newly delivered material it was found that significant

© 2015 American Chemical Society In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch007

hypochlorite degradation takes place. This results in an increase of chlorate and perchlorate levels in the tanks. Dilution of high-strength hypochlorite solutions with softened water and low-strength hypochlorite was examined. The two dilution sources produced similar levels of hypochlorite degradation and chlorate formation. The results were compared to American Water Works Association’s (AWWA) Hypochlorite Assessment Model, which predicts hypochlorite degradation and chlorate production during storage at a constant temperature. Most of the data in the dilution experiments was within 10% deviation from the values predicted by the Hypochlorite Assessment Model. This study also confirmed the results from previous work that in addition to decreasing hypochlorite degradation, dilution also minimizes formation of chlorate and perchlorate.

Introduction Many drinking water treatment facilities have recently switched to hypochlorite as an alternative to chlorine gas due to the dangers associated with the use of chlorine gas. Both chlorine gas and sodium hypochlorite, when added to water, result in the formation of free available chlorine (FAC), primarily in the form of hypochlorite ion or hypochlorous acid (1). FAC is usually expressed in the units of chlorine gas. Some facilities chose to install an onsite hypochlorite generation system (OSG), which typically produces 0.8% FAC hypochlorite solutions. Others use a high concentration hypochlorite provided by an external supplier, which is typically 10-12.5% FAC. It is likely that many more facilities will consider the transition to hypochlorite. However, potential problems associated with hypochlorite storage and use, such as chlorate and perchlorate formation, and excessive chlorine decay, should be considered in the decision process. The focus of the current study is on the high concentration hypochlorite solutions. Hypochlorite manufacture and quality can vary. Hypochlorite also degrades during the storage, thus reducing the chlorine concentration and producing chlorate (2) and perchlorate (3). The loss of chlorine means that more of the solution needs to be applied to achieve the desirable chlorine residual. Therefore, longer periods of hypochlorite storage will lead to higher levels of chlorate and perchlorate in the drinking water. Higher temperatures and direct sunlight contribute to faster hypochlorite degradation. Bromate can also be found in hypochlorite solutions. The main source of bromate in hypochlorite solutions is bromide, which is oxidized during the hypochlorite manufacturing process. Bromide can be present either in the salt or in the water used for the manufacture. The bromate concentration in hypochlorite can be reduced mainly by use of high quality salt and water for hypochlorite manufacture (4). Contrary to chlorate and perchlorate formation, bromate concentrations in hypochlorite do not increase during storage. However, the bromate concentration in the drinking water can increase as a result of 156 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

hypochlorite degradation since more of the solution will be added to the water. Bromate was not the focus of the current study because it does not increase during storage.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch007

Health and Regulatory Aspects Perchlorate is an endocrine disruptor that inhibits iodide intake by the thyroid, thus reducing the production of essential thyroid hormones that are also very important for neurodevelopment (5). Perchlorate is a common contaminant in food and water. It can come from different sources, such as rocket fuel, fireworks and from Chilean nitrate used for agricultural purposes, where it occurs naturally (6). However, the use of hypochlorite as a disinfectant has a potential of significantly increasing perchlorate levels in drinking water, depending on its concentration in hypochlorite solution. Even though perchlorate is not yet regulated on the federal level, the most recent Environmental Protection Agency (EPA) reference dose is established at 0.0007 mg/kg/day. It would imply a drinking water equivalent of 15 ppb, calculated for an average weight of an adult person, 70 kg (7). The states of Massachusetts and California have established a regulation at the levels of 2 ppb and 6 ppb, respectively, based on either an older reference dose, or because they chose to make adjustments to a different primary vulnerable population. Chlorate is used in agriculture as an herbicide and as a bleaching agent in the textile and paper industry. Chlorate is also a disinfection by-product that is introduced into the water during hypochlorite and chlorine dioxide use (8). Intake of high levels of chlorate has resulted in kidney failure and hemolysis. Animal studies prove that chronic and sub-chronic exposure to chlorate has an adverse effect on blood and thyroid (9). A study performed in Italy showed that women exposed to chlorate levels exceeding 200 ppb in drinking water had an elevated risk of having newborns with obstructive urinary defects, cleft palate and spina bifida (10). Chlorate is included in the Third EPA Contaminant Candidate List (CCL3), and has a health reference level (HRL) of 210 ppb (8). Though chlorate is not yet regulated at the national level, the state of California established a notification level of 800 ppb, while the proposed action level is 200 ppb in drinking water. In addition, the World Health Organization (WHO) suggests a guideline of 700 ppb (11). Hypochlorite Assessment Model Several studies that were performed in the last two decades allowed the development of a hypochlorite decay model. The model was developed using mainly laboratory data, from which kinetics of the reactions were inferred. Initially, the model was developed to express hypochlorite decomposition and chlorate formation (2). The agreement between the predicted and measured hypochlorite and chlorate concentrations in the tested solutions had an error of less than 10 percent (2). The model is based on chemical equations that incorporate temperature and ionic strength. Kinetic information was used to develop chemical 157 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch007

and thermodynamic equations that show hypochlorite break-down in pH 11-14 range (12). The model describes the effect of the temperature and ionic strength on the decomposition of OCl− , with the following rate constant relationship (2):

The observed rate constant (k2) includes two independent decomposition pathways of hypochlorite, chlorate pathway and oxygen pathway. Chlorate formation is a result of a slow reaction of two hypochlorite ions followed by a relatively fast reaction of the hypochlorite ion with chlorite:

Therefore, chlorate formation is a second order reaction:

Another pathway, which is however very slow, is the breakdown of hypochlorite to chloride and oxygen, without production of chlorate:

Later, the model was extended to predict perchlorate formation (3). The rate law used is as follows:

The agreement between the observed and measured perchlorate concentrations was ±10% or better. Based on these studies, the model software was developed and currently is available as a tool in the American Water Works Association (AWWA) website under the name of Hypochlorite Assessment Model. The model incorporates the prediction of hypochlorite degradation, and chlorate and perchlorate formation during hypochlorite storage (13). 158 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Objectives The objectives of this study were: 1.

2.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch007

3.

To measure hypochlorite decomposition, and chlorate and perchlorate formation in local systems that are using high-strength hypochlorite in water treatment in the Dallas-Fort Worth (DFW) area, Texas. Compare chlorate and perchlorate levels in these local systems against existing and proposed regulations. Test for possible practical ways to decrease the generation of chlorate and perchlorate, such as dilution of high-strength hypochlorite with softened water produced by an industrial softening system and low-strength on-site generated hypochlorite.

Sample Pretreatment and Analysis Samples were measured for chlorate and perchlorate using ion chromatography (IC) with conductivity detection. This measurement method required pre-treatment of the sample before the analysis. The chlorine needed to be removed prior to the measurement as it is damaging to IC columns. Chlorine removal was achieved by addition of about 20% excess hydrogen peroxide, based on a molar ratio of 1:1 hydrogen peroxide to chlorine. Hydrogen peroxide is also detrimental for IC equipment, therefore it was subsequently removed using manganese dioxide (14). The process showed consistent and reproducible results. Hypochlorite samples were spiked with perchlorate and chlorate standards to check the recoveries of the process. The results of the tests for recoveries of the spiked samples are presented in Table 1. The samples were spiked with three different concentrations, in triplicate for each sample. As can be seen in the table, spike recoveries were above 90%. Calibration curves had correlation coefficients (R2) above 0.99. A more detailed explanation of the pre-treatment method is described in previous work (15). Chlorine levels in hypochlorite were measured using the Hach® DPD colorimetric test (16) and high dilution of hypochlorite or Hach® digital titrator with hypochlorite titration kit (17).

Field Sampling and Results Description and Methodology To assess the first two objectives of the study, there was a need to collect real data from water treatment facilities. One-time sampling from two water treatment facilities was performed in August 2014. Both facilities operate with a residual volume in hypochlorite storage tanks. Hence, tanks are not drained in between the deliveries, but newly delivered material is mixed with the tank content. At each facility, a sample from the storage tank and a delivery truck was collected. Hypochlorite samples were collected into high-density polyethylene (HDPE) bottles and then directly taken to the laboratory using an ice chest 159 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch007

filled with ice. In the laboratory, they were stored in a refrigerator at 4 ºC. Samples were analyzed for free chlorine, chlorate and perchlorate. Chlorine was measured using digital titrator and hypochlorite titration kit. Theoretical chlorate and perchlorate concentrations in drinking water were calculated based on the hypochlorite strength and application of 5 mg-Cl2/L, which is the average application dose at the facilities (see eq. 8). Both facilities apply hypochlorite along with ammonia to generate monochloramine as their residual secondary disinfectant. A dose of 5 mg-Cl2/L of hypochlorite along with a corresponding amount of ammonia typically results in a residual monochloramine concentration of between 3.5 and 4.0 mg-Cl2/L at both plants. The chlorate and perchlorate concentrations in the finished water are tied to the hypochlorite dose and not the residual monochloramine concentration.

Table 1. Spike Recoveries Data for Perchlorate and Chlorate Perchlorate (spike of freshly delivered sample-concentrations in ppm) not spiked

spiked 1

spiked 2

spiked 3

Not detectable

4.902

9.617

23.37

SD

0.081

0.496

0.111

Average % recovery

98.0

96.2

93.5

%RSD

1.66

5.16

0.48

Average (n=3)

Chlorate (spike of diluted freshly delivered sample-concentrations in ppm) Average (n=3)

35.60

38.48

41.25

51.24

SD

0.268

0.274

1.152

1.837

93.1

92.1

98.1

0.71

2.79

3.59

Average % recovery %RSD

0.75

It should be noted that in this experiment the samples were not tested in replicates for chlorate and perchlorate, therefore standard deviations were not calculated. However, based on the previous measurements, average relative standard deviation (RSD) is 2.4% for perchlorate and 2.0% for chlorate. (calculated from values demonstrated in Table 1). Results from the water treatment facilities are presented in Table 2. Expected chlorate concentration was calculated as follows:

Similarly, the expected perchlorate concentration in finished water was calculated. 160 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Results

Table 2. Chlorine Degradation and Disinfection By-Products Formation in Hypochlorite Storage Tanks in the Month of August Facility

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch007

Facility I

Facility II

Description

Chlorine, as Cl2

Chlorate, as ClO3–

Perchlorate, as ClO4–

Newly-delivered

108 g/L

947 ppm

94%

Concentration in water assuming 5 mg/L chlorine application dose

N/A

338 ppba

0.025 ppbb

Newly-delivered

133 g/L

1315 ppm

1195%

Concentration in water assuming 5 mg/L chlorine application dose

N/A

467 ppba

0.153 ppbb

Health Reference Level for chlorate: 210 ppb. b Currently regulated at MCL of 2 ppb and 6 ppb in Massachusetts and California, respectively.

a

At Facility I, the chlorine level in the newly delivered material was 108 g-Cl2/L. However, the chlorine level in the tank was lower, 96 g-Cl2/L. This means that there is a hypochlorite degradation as a result of the storage in the hypochlorite tank. In this specific case, the difference in chlorine concentrations between new and stored hypochlorite was 11%. At Facility II, the new material had chlorine concentration of 133 g-Cl2/L, while the stored material was 106 g-Cl2/L. In this case, the difference was 21%. A possible reason for this variation in difference between new and stored material at both plants could be the difference in the strength of the delivered hypochlorite, which is 10% at Facility I and 12.5% at Facility II. According to eq. 4, hypochlorite degradation is a second order reaction on hypochlorite ion, therefore for higher hypochlorite concentration the degradation rate will be higher. Other possible reasons include variation in operational parameters, such as storage time, volume of the tank and level of the solution in the tank. In addition to hypochlorite degradation, chlorate and perchlorate formation takes place. Chlorate and perchlorate levels in the storage tanks were much higher than in new delivered material at both plants (Table 2). To access chlorate concentrations in finished water treated by the facilities, eq. 8 was used. Expected perchlorate levels in water were calculated in a similar way. 161 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch007

Based on the results from the tank samples from August and assuming a chlorine dose of 5 mg-Cl2/L, chlorate concentrations in the tanks are high enough that it would probably cause the concentrations in water to exceed the health reference level, which is 210 ppb. Despite perchlorate concentrations rising during the storage, the residual concentrations in finished water would probably be below the typical standards (lowest standard is 2 ppb) and therefore seem to be not a concern. August temperatures in DFW often reach triple-digit values, which accelerates hypochlorite degradation and chlorate and perchlorate formation in storage tanks. Two methods to reduce hypochlorite degradation and chlorate and perchlorate formation are: 1. 2.

Increasing the frequency of deliveries and/or reducing the residual storage volume, thus increasing the hypochlorite turnover rate, and Diluting high-strength hypochlorite.

Investigation of Dilution Effects Previous studies indicated that one of the possible ways of reducing the formation of chlorate and perchlorate is dilution of high-strength hypochlorite solution (4). In the current study, high-strength hypochlorite was diluted with softened water and low-strength hypochlorite at various ratios and the impact on hypochlorite degradation and chlorate and perchlorate formation was tested. Description and Methodology Two practical ways for diluting high-strength hypochlorite were tested and then compared to the Hypochlorite Assessment Model: dilution with softened water and dilution with low-strength hypochlorite. Both solutions used for dilution were collected from an OSG water treatment plant. While it is more likely that softened water will be used for dilution, low-strength hypochlorite was also tested to examine whether it yields better results in terms of slower hypochlorite degradation and chlorate formation. The use of low-strength hypochlorite for dilution is possible only for plants that produce OSG hypochlorite, but desire to store bulk hypochlorite for emergency purposes. Therefore, the intention was to check the possibility of a bulk hypochlorite dilution with OSG hypochlorite for such a facility. A sample of high-strength hypochlorite was collected from a delivery truck at one of the facilities that was a part of this study. For the purposes of dilution, low-strength hypochlorite (0.8%) and softened water samples were collected from a facility generating on-site low-strength hypochlorite (0.8%). Hypochlorite samples at different dilution ratios were incubated to check the impact of dilution on the stored material. Five samples at different dilution ratios were incubated at temperature of 40 ºC for 63 days. Temperature of 40 ºC was considered to account for the worst case scenario. August temperatures in 2014 in DFW reached 102 ºF, 39 ºC (18). 162 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

• • •

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch007

• •

Sample 1: 100% High-strength hypochlorite Sample 2: 50% High-strength hypochlorite + 50% Softened water Sample 3: 50% High-strength hypochlorite + 50% Low-strength hypochlorite Sample 4: 25% High-strength hypochlorite + 75% Softened water Sample 5: 25% High-strength hypochlorite + 75% Low-strength hypochlorite

Samples were placed into 2 L HDPE bottles, and every 7 days a 100 ml sample was collected from each bottle. All the samples were analyzed for chlorine and chlorate. High-strength hypochlorite and one of the samples diluted with 1:1 ratio were analyzed for perchlorate. It should be noted that in this experiment the samples were not analyzed in triplicate for chlorate and perchlorate due to the vast number of samples and laboratory limitations, and because the primary goal of this experiment was to determine the trends rather than exact concentrations. Due to this reason, standard deviations could not be calculated and added to the graphs. Total chlorine measurements were done in triplicate with the DPD Hach® kit. The use of the DPD kit required very high dilutions (20,000:1), which can introduce an error into the results, especially when dealing with low concentrations. This could be a possible reason for several points with deviations from the trend, especially in the four-fold dilution. Results from various dilution experiments are presented in Figures 1-3. Results The degradation of hypochlorite in the incubated solutions is shown in Figure 1. It can be seen that the undiluted sample starts with relatively high chlorine concentration, 112 g-Cl2/L and after 60 days degrades to about 1/3 of its initial chlorine concentration, 36 g/L. The degradation is much slower for solutions that were diluted with 1:1 dilution ratio and slows down even more for solutions with 3:1 dilution ratio. Hence, the decrease in chlorine strength will be higher for higher strength hypochlorite solutions, a fact that corresponds to the previous studies (4). Also, no significant difference can be seen in samples diluted with low-strength hypochlorite (0.8%) and samples diluted with softened water at the same ratios. Similarly to hypochlorite degradation, chlorate formation is much higher in high-strength solutions (Figure 2). There was an about 30 fold increase in the chlorate concentration in the high-strength solution relative to the initial chlorate concentration (from 1.256 g/L to 28.151 g/L). Chlorate formation slows down significantly at 1:1 dilution ratio, and even more at 3:1 dilution. Samples diluted with low-strength hypochlorite (0.8%) and samples diluted with softened water at the same ratios had similar chlorate formation trends. The formation of perchlorate in the high-strength sample was compared to a sample diluted with the softened water at a 1:1 dilution ratio (Figure 3). Perchlorate comparison was limited to these two samples due to laboratory constraints. However, these two samples are sufficient to show that there is a significant decrease in perchlorate formation even at 1:1 dilution ratio. 163 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch007

Figure 1. Hypochlorite solutions degradation as a function of time and initial concentration at 40 ºC.

Figure 2. Chlorate formation in hypochlorite solutions as a function of time and initial concentration at 40 ºC.

Dilution decreases hypochlorite decomposition and oxyhalide formation by decreasing hypochlorite ion concentration and by decreasing ionic strength, which both contribute to hypochlorite decomposition (2). It was shown in previous work that a two-fold dilution slows down perchlorate formation by the factor of 7 (4), 164 In Trace Materials in Air, Soil, and Water; Evans, Kendra R., et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Publication Date (Web): December 1, 2015 | doi: 10.1021/bk-2015-1210.ch007

and chlorate formation by the factor of 5 (2). Higher dilution ratios will reduce hypochlorite degradation to a higher extent (6), however, it is recommended that the pH of the diluted solution will be kept between 12 to 13 (19). Following four fold dilution, the pH of the tested samples was 12.12, given that the initial pH of bulk sample was 12.75, as shown in Table 3. In case of lower pH of the original sample, four-fold dilution is more likely to result in pH