Identification of inhalable rutile and polycyclic aromatic hydrocarbons (PAHs) nanoparticles in the atmospheric dust

Journal Pre-proof Identification of inhalable rutile and polycyclic aromatic hydrocarbons (PAHs) nanoparticles in the atmospheric dust Ana L. Gallego-Hernández, Diana Meza-Figueroa, Judith Tanori, Mónica AcostaElías, Belem González-Grijalva, Juan F. Maldonado-Escalante, Sarai Rochín-Wong, Diego Soto-Puebla, Sofia Navarro-Espinoza, Roberto Ochoa-Contreras, Martín Pedroza-Montero PII:

S0269-7491(19)33078-7

DOI:

https://doi.org/10.1016/j.envpol.2020.114006

Reference:

ENPO 114006

To appear in:

Environmental Pollution

Received Date: 11 June 2019 Revised Date:

13 January 2020

Accepted Date: 15 January 2020

Please cite this article as: Gallego-Hernández, A.L., Meza-Figueroa, D., Tanori, J., Acosta-Elías, Mó., González-Grijalva, B., Maldonado-Escalante, J.F., Rochín-Wong, S., Soto-Puebla, D., NavarroEspinoza, S., Ochoa-Contreras, R., Pedroza-Montero, Martí., Identification of inhalable rutile and polycyclic aromatic hydrocarbons (PAHs) nanoparticles in the atmospheric dust, Environmental Pollution (2020), doi: https://doi.org/10.1016/j.envpol.2020.114006. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2020 Published by Elsevier Ltd.

Rutile (TiO2) and PAHs NPs inhalation

Resuspension . . . . . by wind, . . . . . . . . . . . vehicle and . . . . . . . . . . . . . . . . . pedestrian . . . . . . . . . traffic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Health complications: . . . •  Respiratory diseases . . . •  Cardiovascular dysfunction . . . . . . . . . . . •  Cancer . . . . •  Cyto and genotoxicity

1

Identification of inhalable rutile and polycyclic aromatic hydrocarbons (PAHs)

2

nanoparticles in the atmospheric dust

3

Ana L. Gallego-Hernández,a Diana Meza-Figueroa,b Judith Tanori,c Mónica Acosta-

4

Elías,a Belem González-Grijalva,d Juan F. Maldonado-Escalante,a Sarai Rochín-Wong,c

5

Diego Soto-Puebla,a Sofia Navarro-Espinoza,e Roberto Ochoa-Contreras,f and Martín

6

Pedroza-Monteroa *

7

a

8

Sonora, México.

9

b

10

c

11

Hermosillo 83000, Sonora, México.

12

d

13

Autónoma de México

14

e

15

Hermosillo 83000, México.

16

f

17

México.

18

*Corresponding author:

19

Martín Pedroza-Montero, PhD

20

Address: Rosales y Encinas. Hermosillo, Sonora, México. 83000

21

Phone: +52-(662)-259-2156

22

[email protected]

Departamento de Investigación en Física, Universidad de Sonora, Hermosillo 83000,

Departamento de Geología, Universidad de Sonora, Hermosillo 83000, Sonora, México.

Departamento de Investigación en Polímeros y Materiales, Universidad de Sonora,

Posgrado en Ciencias de la Tierra, Instituto de Geología, Universidad Nacional

Posgrado en Nanotecnología, Departamento de Física, Universidad de Sonora,

Centro de Investigación en Alimentación y Desarrollo, Hermosillo, Sonora, 83304,

23

Abstract

24

Addressing the presence of rutile nanoparticles (NPs) in the air is a work in progress, and

25

the development of methodologies for the identification of NPs in atmospheric dust is

26

essential for the assessment of its toxicological effects. To address this issue, we selected

27

the fast growing desertic city of Hermosillo in northern Mexico. Road dust (n=266) and

28

soils (n=10) were sampled and bulk Ti-contents were tested by portable X-ray

29

fluorescence. NPs were extracted from atmospheric dust by PM1.0-PTFE filters and

30

further characterized by Confocal Raman Microscopy, Energy-dispersive X-ray

31

spectroscopy (EDS) coupled to Transmission Electron Microscopy (TEM) and Scanning

32

Electron Microscopy (SEM). Results showed (i) the average concentration of Ti in road

33

dust (3447 mg.kg-1) was similar to natural values and worldwide urban dusts; (ii) the bulk

34

geochemistry was not satisfactory for Ti-NPs identification; (iii) 76% of the total

35

extracted PM1.0 sample corresponded to NPs; (iv) mono-microaggregates of rutile NPs

36

were identified; (v) ubiquitous polycyclic aromatic hydrocarbons (PAHs) were linked to

37

NPs. The genotoxicity of rutile and PAHs, in connection with NPs content, make us

38

aware of a crucial emerging environmental issue of significant health concern, justifying

39

further research in this field.

40 41

Capsule:

42

Characterization of airborne NPs will help us identify the potential risks to human health

43

and ecosystems.

44

Keywords:

45

Atmospheric dust; PM1.0; arid areas

46

1. Introduction

47

Air pollution is a heterogeneous mixture of gases, biomolecules, and particles with an

48

aerodynamic size smaller than 100 µm (Andreau et al., 2012; Slezakova et al., 2013). In

49

2018, the World Health Organization (WHO) estimated that 4.2 million deaths per year

50

are due to ambient air pollution and reported that 91% of the world’s population lives in

51

places with low air quality. The air pollutants with the highest health concern are ozone

52

(O3), nitrogen dioxide (NO2), and sulfur dioxide (SO2) linked to combustion processes, as

53

well as particulate matter (PM) from exhaust and non-exhaust sources (Andreau et al.,

54

2012; Franchini and Mannucci, 2007; Slezakova et al., 2013). Airborne PM is

55

categorized according to its aerodynamic diameter as coarse (<10 µm), fine (<2.5 µm),

56

and ultrafine or nanoparticles (<100 nm) (Franchini and Mannucci, 2007; Nagar et al.,

57

2014). The size of the particles has been directly linked to their potential to cause health

58

problems. Fine and ultrafine PM are of particular interest due to their ability to penetrate

59

the tracheobronchial and alveolar regions of the lungs after inhalation and reaching other

60

organs via the bloodstream (Krug and Wick, 2011; Oberdörster, 2001; Oberdörster et al.,

61

2005; Zou et al., 2017). They have been associated with respiratory diseases such as lung

62

cancer, chronic obstructive pulmonary disease and asthma, as well as cardiovascular

63

disorders, and premature death (Chehregani et al., 2004; Evans et al., 2014; Jakubiak-

64

Lasocka et al., 2015; Jiménez et al., 2010; Jin et al., 2018; Kim et al., 2015; Longhin et

65

al., 2018, 2016, 2013; Lu et al., 2014; Pope et al., 2004; Ribeiro et al., 2015; Sedghy et

66

al., 2018).

67

The contribution of nanoparticles (NPs) to the total mass concentration of PM in air

68

pollution is very low (Calderón-Garcidueñas et al., 2019). However, they constitute the

69

majority of particle numbers that are unregulated (Rönkkö et al., 2017). In atmospheric

70

urban environments, NPs can be divided into two groups: (i) Primary particles, derived

71

from discharges from construction areas, erosion of pavement road-dust (friction), and

72

combustion (Silva et al., 2020); and (ii) Secondary particles which are formed in the air

73

by the geochemical interaction of primary particles and the air gaseous pollutants from

74

industries and vehicular traffic (Morillas et al., 2018a and 2018b).

75

The behavior of natural nanoparticles (NNPs) has been widely studied, but the

76

identification of NPs in real-world atmospheric environments is limited, precluding the

77

evaluation of potential exposure to humans (Silva et al., 2020; Wagner et al., 2014).

78

Common examples of engineered nanoparticles (ENPs) include fullerene, silver, zinc

79

oxide, iron oxide, quantum dots, single-walled carbon nanotubes, multiwalled carbon

80

nanotubes (Navarro et al., 2008). Titanium dioxide ENPs (TiO2) are one of the most

81

common produced nano-materials worldwide (>10,000 t/a), despite their reported adverse

82

impacts on human health (Oberdörster, 2001; Piccinno et al., 2012; Saquib et al., 2012).

83

TiO2 NPs have been widely used in construction materials to improve the mechanical

84

performance of concrete, ceramic, steel or paints, and as an additive in many foods,

85

personal care items and other consumer products (Lee et al., 2010; Oliveira et al., 2019;

86

Piccinno et al., 2012; Robichaud et al., 2009; Weir et al., 2012). Rutile is the most

87

common form of TiO2 (Meinhold, 2010). Due to its high specific gravity, high refractive

88

index and hardness, rutile is being increasingly used as NPs in a multitude of products

89

(Farjana et al., 2018; Gázquez et al., 2014). Rutile enhances the performance of metal

90

parts in aircraft engines, sporting equipment, and pigments. Rutile is widely used as a

91

whitener, in paints, adhesives, plastics, ceramics, paper, sunscreens, food as well as

92

several additional applications in nanotechnology.

93

Rutile may have an either natural or anthropic origin. This mineral is an accessory in

94

soils, and it can be found as NPs associated with heteroaggregates of Fe, Mn, Si and Al

95

particles (Schindler and Hochella, 2016). The morphology of TiO2 NPs is suggested as a

96

potential criterion to distinguish among natural vs anthropic rutile (Pradas del Real et al.,

97

2018). Other authors suggest the presence of engineered “organic coating” for ENPs as

98

an indicator of origin, but the organic compounds of the layer covering the ENPs could

99

be easily desorbed by water-interaction in nature (Wagner et al., 2014). The removal of

100

such coating is feasible because ENPs of similar composition as NNPs can follow equal

101

transformation pathways in the environment, thus avoiding the distinction between

102

sources. Most published studies on NNPs are limited to aquatic environments (Wu et al.,

103

2020; Hartland et al., 2013) with scarce studies in the atmosphere (Silva et al., 2020). The

104

behavior of NNPs in water bodies and soils has been described as chemically reactive and

105

mobile (Sebesta et al., 2020; Loosli et al., 2019; Hartland et al., 2013). In natural systems,

106

NPs can form aggregates with products of biological decay such as humic matter, and

107

minerals produced by the chemical weathering of rocks (oxides, oxyhydroxides of iron,

108

manganese and aluminum, as well as aluminosilicates (Hartland et al., 2013). The

109

mobility of NPs in natural environments is variable. In surface waters, the mobility of

110

NPs is related to their colloidal stability, whereas in fractured aquifers NPs move freely.

111

In alluvial groundwater aquifers and soils, the movement of NPs is constrained by the

112

potential collision with soil grains (Cullen et al., 2010; Nikolla, 2008). The presence of

113

NPs in the atmospheric environment results from liquid condensation and interactions of

114

gaseous/solid PM (Ribeiro et al., 2010).

115

ENPs result in rapid aggregation because of surface charge interactions due to the

116

engineered coating, thus forming mono-microaggregates with (i) organic matter as

117

reported in sludges (Pradas del Real et al., 2018) or (ii) Fe-spheres in combustion

118

products (Calderón-Garcidueñas et al., 2019).

119

Despite their origin, potentially harmful consequences of NPs for human health makes

120

crucial the development of methodologies to facilitate the detection of their occurrence in

121

environmental samples (Bundschuh et al., 2018; Von Der Kammer et al., 2012). In this

122

work, airborne urban dust in the size fraction of 1.0 µm or less (PM1.0) was evaluated in

123

relation to their size and chemical composition to determine of NPs of rutile and PAHs

124

are in the airborne respirable fraction. Hermosillo city was chosen as the study site

125

because it is located within the Arizona-Sonora desert in northern Mexico, and because

126

arid conditions promote dust emissions and the transport of PM. The aim of this study

127

was 1) to extract crystalline NPs from atmospheric dust; and 2) to identify the presence of

128

potential anthropogenic rutile NPs in the atmospheric environment.

129

130

2. Materials and methods

131

2.1 Study area

132

Hermosillo is located in the Sonoran Desert of northern Mexico. The population is nearly

133

900,000 inhabitants. In the last few years, the city has experienced significant growth,

134

thus impacting the vehicular traffic volume, as well as urbanized areas where

135

construction activities have increased (COESPO, 2015). The climate is dry for most of

136

the year, and the region is affected by both brief and intense rainfall during the

137

summertime. Dust emissions are common in the area due to the erosive potential of the

138

North American Monsoon. The lack of effective rainwater drainage systems in

139

Hermosillo causes strong surface run-off. Previous studies showed that traffic sources

140

may increase Ti content in atmospheric dust collected at pedestrian levels (Meza-

141

Figueroa et al., 2016). The erosion of the constructed urbanized area and the effect of

142

traffic enhances suspension processes (Meza-Figueroa et al., 2016; Moreno-Rodríguez et

143

al., 2015).

144 145

2.2 Road dust sample collection, and preparation

146

Road dust samples were collected from 226 locations evenly distributed within the city,

147

and 10 more sites outside city limits were used to determine the local geochemical

148

background (LGB) (Figure 1A). Sample collection was performed as indicated in Meza-

149

Figueroa et al. (2018). Briefly, samples were sieved, and the final fractions that passed

150

through the #635 mesh were obtained. These fractions correspond to a particle size

151

smaller than 20 µm in aerodynamic diameter.

152 153

2.3 PM sampling site, collection, and extraction

154

For atmospheric dust collection, the sampling site was selected on the intersection of two

155

high traffic roads in the geographic center of Hermosillo, Sonora, Mexico. A frmOMNItm

156

air sampler (BGI, USA) with a 5 L/min miniPM inlet for PM1.0 was placed at the

157

pedestrian level to collect resuspended dust from traffic activity. These collectors had

158

previously been used in environmental studies (Sharma et al., 2013). The air sampler was

159

operated for 10 hours set to collect dust using a PTFE Filter with a PFA support ring

160

(MTL #PT47AN). Standard protocols have been established for NPs extraction in

161

biological applications (Perrone et al., 2010), but, to our best knowledge, extraction

162

procedures for NPs in air filters are not documented. We found a few references about the

163

presence of NPs in natural environmental samples (Tong et al. 2015; Oliveira et al. 2019;

164

Pradas del Real et al., 2019) but in these works, NPs were extracted from sludges, water

165

and construction-derived dust. In this research, we modified a protocol for NPs extraction

166

from PM1.0 PTFE air filters. Glass fiber filters are commonly used in air quality

167

monitoring, but NPs extraction is extremely difficult due to the release of insoluble

168

microfibers during the ultrasonication. For particle extraction, PFA rings were carefully

169

removed with a clean and sterile blade, and the filters were placed in a glass vial with 3

170

ml of Milli-Q water (Perrone et al., 2010). Air particles were extracted from the filter by

171

ultrasonication in a Branson Ultrasonic M2800 for 20 minutes and transferred to a glass

172

Petri dish for drying in a desiccator. Particles were resuspended in 1 ml of ultrapure

173

water, and the concentration calculated according to the weight.

174 175

2.4 X-Ray Fluorescence analysis of road dust samples

176

A portable NITON FXL X-Ray Fluorescence (XRF) analyzer (Thermo Scientific) was

177

used to obtain the elemental concentration (Meza-Figueroa et al., 2018). The

178

determination of elemental concentration was evaluated by following the procedures

179

described in Method 6200 of the United States Environmental Protection Agency

180

(USEPA, 2007) using portable X-ray fluorescence equipment. Silicon dioxide (quartz)

181

free of any analyte at concentrations above the established lower limit of detection was

182

used as blank. For standard reference materials (SRMs), containing certified amounts of

183

metals in soil or sediment, the NIST standard reference materials 2710, and 2711 were

184

used. The precision (%RSD) and the accuracy (%) were evaluated based on the analysis

185

of seven replicates. An acceptable %RSD obtained range is 80 to 120 for analyzed

186

elements. Detection limits are expressed in mg.kg-1. A map with road dust Ti

187

concentrations was generated with the geographic information system ArcGIS. The

188

Minitab 17 software (version 17.1.0.0, 2013, Minitab Inc., State College, PA, US) was

189

used for statistical analysis.

190 191

2.5 Scanning Electron Microscopy coupled to Energy-dispersive X-ray spectroscopy

192

(SEM-EDS)

193

The Phenom-Pro X Scanning Electron Microscope (SEM) with an Energy-Dispersive X-

194

Ray Spectrometer (EDS) was used for imaging and elemental analysis on the selected

195

PM samples. Extracted PM were collected from the Petri dish with the SEM pin holder.

196

Particles as agglomerates directly from air filters were studied with a Hitachi TM3030

197

benchtop SEM and a Bruker Quantax 70 Energy-Dispersive X-Ray Spectroscopy at the

198

National Laboratory of Geochemistry and Mineralogy.

199 200

2.6

201

spectroscopy (TEM-EDS)

202

PM were collected with a 400 mesh copper grid directly from the PTFE filter by scraping

203

the surface of the filter several times and letting particles adhere to the grid. The

204

morphology and size distribution of selected samples were determined by using a JEM

205

2010F Transmission Electron Microscope (TEM) with an operating voltage of 200 kV

206

(JEOL, Ltd., Tokyo, Japan). All samples were subsequently dried in a vacuum before

207

observation. The characteristic chemical elements were identified by EDS analysis

208

(Quantax 200 X-ray energy dispersive spectrometer, Bruker, GmbH, Berlin, Germany).

209

The TEM-EDS analysis was carried out with two independent collected samples, at the

210

TEM Laboratory of the University of Sonora. TEM images of total PM were analyzed

211

using Image J to determine particle size distribution. For High Resolution-Transmission

212

Electron Microscopy (HR-TEM) analysis, images were processed with the Gatan

213

Microscopy Suite Software to calculate the interplanar distance of the crystals.

214

Mineralogical identification of rutile was not conducted by powder X-ray diffraction due

215

to the small number of samples. Instead, the Diffraction Files from Bruker Diffrac.eva

216

plus software were used.

Transmission

Electronic Microscopy coupled

to

Energy-dispersive X-ray

217 218

2.7 Confocal Raman Microscopy

219

The Raman spectra of filters were acquired with a Raman microspectrometer (Witec,

220

alpha300 RA, Ulm, Germany) using a frequency-doubled Nd:YAG laser excitation of

221

532 nm (CW–continuous wave), with a 600 gr/mm grating. For Raman spectroscopy

222

analysis, 40 μl of the PM sample resuspended in highly pure water was placed on a

223

calcium fluoride (CaF2) slide and dried in a desiccator for 2 hours. Adhered particles

224

within an area of 50 × 50 µm were analyzed, using an integration time of 1 s/spectrum, 10

225

mW of laser power, and an objective 100X.

226 227

2.8 Geochemical indices

228

To estimate the level of titanium contamination in the studied samples, the (i) enrichment

229

factor, (ii) contamination factor, and (iii) geoaccumulation index were calculated. These

230

indices are commonly used in urbanized areas to assess post-industrial pollution (Kusin et

231

al., 2019). Titanium enrichment was calculated using enrichment factor (EF) after Buat-

232

Menerd and Chesselt (1979)

233 234

EF= (Ti/Fe)sample/ (Ti/Fe)background

235 236

where Fe is the reference element. Background values were obtained from the average of

237

10 superficial soil samples taken from periurban sites in Hermosillo city. Classification of

238

EF is described after Sutherland (2000) as follows: EF<1 no enrichment; EF <3 minor

239

enrichment; 3 ≤ EF < 5 moderately enrichment; 5 ≤ EF < 10 moderately to severe

240

enrichment; 10 ≤ EF < 25 severe enrichment; 25 ≤ EF < 50 very severe enrichment; EF ≥

241

50 extremely enrichment.

242

The contamination factor (CF) is useful to estimate the contamination status by a single

243

substance. The calculation used Hakanson (1980) as follows

244

245

CF= sample/background

246 247

where sample is the measured content of Ti in the samples and background is the value of

248

Ti in uncontaminated local geochemical background. Contamination factor is classified

249

as following: CF < 1 low contamination factor; 1 ≤ CF < 3 moderate contamination

250

factor; 3 ≤ CF < 6 considerable contamination factor; CF ≥ 6 very high contamination

251

factor.

252 253

The geoaccumulation index has been described by Muller (1969) as:

254 255

Igeo= log2[m/(1.5xbackground)]

256 257

where m is the concentration of Ti in the sample, and background is the concentration of

258

Ti in the local geochemical background. Degree of pollution is classified as follows:

259

Igeo<0 uncontaminated; 0 ≤ Igeo < 1 uncontaminated to moderately contaminated; 1 ≤

260

Igeo < 2 moderately contaminated; 2 ≤ Igeo < 3 moderately to strongly contaminated; 3 ≤

261

Igeo < 4 strongly contaminated; 4 ≤ Igeo < 5 strongly to extremely contaminated; Igeo≥ 6

262

extremely contaminated.

263 264

3. Results

265

3.1 Ti concentration in road dust and PM

266

To obtain a first indication of the Ti spatial distribution in road dust, total Ti content was

267

analyzed to identify the accumulation areas in Hermosillo, Mexico. These sites represent

268

areas where dust emission could be enhanced. Evenly distributed study sites in the urban

269

area were selected for analysis (n=226). Thus, a study site was chosen close to downtown

270

with high vehicle and pedestrian traffic to obtain a more detailed characterization (Figure

271

1A, white star). Additionally, 10 locations were selected outside city limits to calculate

272

local geochemical background (LGB), from superficial soils non-impacted by

273

anthropogenic activities (non-natural sources).

274

The results showed Ti-concentrations levels ranging from 1248 to 8409 mg.kg-1, with a

275

mean value of 3447 mg.kg-1 for dust collected within the urbanized area (Table 1, Figure

276

1B). Taylor & McLennan (1985) reported a Ti-average continental crust value of 5400

277

mg.kg-1, and Bowen (1979) documented a mean Ti-value of 5000 mg.kg-1 for worldwide

278

soils. Average concentration of Ti in the collected dust samples slightly exceeded LGB

279

but within natural contents ranges (Figure 1B). The study site is representative since Ti-

280

content in bulk dust (4200 mg.kg-1) is similar to the mean Ti-value in the city when

281

considering all 226 collected dust samples (Table 1). Enrichment factor (EF),

282

contamination factor (CF) and geoaccumulation index (Igeo) were calculated for Ti

283

maximum content in the samples. An EF value of 0.42 was obtained indicating no

284

enrichment. On the other hand, a CF of 2.48 showed moderate contamination, and a value

285

for Igeo of 0.72 indicated uncontaminated to moderately contaminated. With these data

286

we can conclude a moderate anthropic contribution of TiO2. According to the obtained

287

indices, bulk geochemistry cannot be considered as an indicator of the presence of TiO2

288

NPs. This is because their contribution to total mass based geochemistry is negligible. Ti-

289

contents in road dust in Hermosillo city have similar values to the ones found in several

290

reports from various other locations around the world, indicating that Ti-NPs may remain

291

undetected elsewhere (Table 1).

292 293

3.2 PM and NPs agglomerates

294

To analyze atmospheric dust at our representative study site, we collected PM at the

295

pedestrian level. The chemical identification of airborne particles was performed by

296

SEM-EDS. Figure 2 shows aggregated particles formed by titanium (Ti) and iron (Fe)

297

similar to those reported by Adachi and Buseck (2010). Transmission Electron

298

Microscopy (TEM) coupled to Energy-Dispersive X-Ray Spectroscopy (EDS) was used

299

to identify the morphology and size of particles (Figure 3A). In this work,

300

monocrystalline aggregates of smooth-shaped TiO2 were identified (Figure 3B). Also, in

301

the extracted fraction, NPs size distribution showed a prevalent size range from 2 nm to

302

650 nm, and the most abundant NPs size being between 41 and 60 nm (Figure 3C).

303

Evidence of sample compositional complexity is provided by the EDS semiquantitative

304

analysis. Results indicated the presence of titanium (Ti), oxygen (O), and silicon (Si)

305

(Figure 3D), as well as calcium (Ca), sodium (Na), potassium (K), sulfur (S), aluminum

306

(Al), magnesium (Mg), chlorine (Cl), zirconium (Zr) and iron (Fe) (data not shown).

307

Peaks of copper (Cu) might be generated by the grid.

308

To further validate the results obtained from TEM analysis, we used SEM-EDS to

309

characterize the extracted particles. As shown in Figure 4, particles composed of O and

310

Ti were present in the sample, suggesting the presence of NPs of titanium dioxide (TiO2).

311

Consistent with TEM-EDS, other elements detected in the sample were silicon (Si),

312

strontium (Sr), zinc (Zn), nitrogen (N), boron (B), molybdenum (Mo), phosphorous (P)

313

and bromine (Br), and indicating that the sample composition was not affected by the

314

extraction method (data not shown). Results obtained by both methodologies depicted

315

heterogeneity of the sample, but Ti-oxides were identified independently of the method

316

and the extraction process.

317

Pradas Del Real et al (2018) proposed the morphology of TiO2 particles as a potential

318

criterion to provide insights into the sources of natural vs anthropic rutile. Rutile as NNPs

319

is characterized by rough irregular shapes whereas ENPs form smooth shaped particles.

320

Wagner et al (2014) developed a decision tree model for the identification of NPs of

321

natural and anthropic origin. For ENPs with a natural counterpart, such as rutile, the

322

presence of engineered coating is crucial to identify the source. However, in the absence

323

of such coating, NNPs and ENPs could be indistinguishable. Soil inorganic NPs are

324

mostly manganese/ iron oxyhydroxides, and clays (Theng and Yuan, 2008). Main

325

crystalline forms of TiO2 are anatase, rutile, and they may precipitate after weathering of

326

kaolinite in tropical soils (Cornu et al., 1999). Few articles describe the presence of rutile

327

NPs in soils (Dias et al, 2013; Taboada et al., 2006).

328

Even though our data are inconclusive for source apportionment of rutile, a natural origin

329

as the exclusive source for rutile in the area seems unlikely due to the local geology and

330

soil-formation processes in arid environments. Rutile is commonly found as a minor

331

phase in high-temperature/high-pressure metamorphic rocks. It is also found as a limited

332

accessory mineral in igneous rocks, mainly granitic. Main lithology in the area is

333

dominantly granite-limestone and a few mafic dykes. The highest reported TiO2 value in

334

studied dust is 8409 mg.kg-1, which is anomalous when compared to average continental

335

crust values of 5400 mg.kg-1 (Taylor & McLennan, 1985), and a mean value of 5000

336

mg.kg-1 for worldwide (Bowen,1979), this shows a possible anthropogenic contribution

337

of TiO2 to urban dust, but it remains unknown if this value is related to rutile, so further

338

studies are recommended.

339 340

3.3 Identification of rutile crystals and polycyclic aromatic hydrocarbons (PAHs)

341

Figure 5 shows (A) optical, and (B) Raman spectroscopy images of the studied sample.

342

Scanning Raman Spectroscopy results showed that the PM1.0 sample consists of a diverse

343

mix of components (Figure 5B). Cluster analysis obtained with WiTec software (Witec,

344

Ulm, Germany) showed four different Raman spectra (Figure 5C and D).

345

Three well-defined peaks at 240, 438 and 603 cm-1 (Figure 5C, in blue), revealed the

346

presence of titanium dioxide (TiO2) in the atmospheric sample (Frank et al., 2012). This

347

result validated the data obtained from the TEM-EDS and SEM-EDS for elemental

348

determination, showing high concentrations of titanium (Ti) and oxygen (O) in the

349

sample. Furthermore, polycyclic aromatic hydrocarbons (PAHs) were also detected, with

350

peaks between 995 and 1562 cm-1 (Figure 5C and D, red) (Chen et al., 2014). These

351

PAHs are widespread atmospheric pollutants, formed by the incomplete combustion of

352

fossil fuels and other carbon-containing fuels with carcinogenic, mutagenic and cytotoxic

353

activities (Agarwal et al., 2018; Billet et al., 2007; Ravindra et al., 2008; Saint-Georges et

354

al., 2007). In particular, it is documented that Fe-rich, magnetic and combustion/friction

355

derived NPs make up almost 10% of the solid NPs reported in atmospheric samples

356

collected at Mexico´s city (Calderón-Garcidueñas et al., 2019). Previous reports have

357

shown that PM-bound PAHs are associated with an increased hazard to human health

358

(Zhang et al., 2019).

359

In arid zones, vehicle exhaust releases PM at roadsides, as ultrafine particles, and may

360

form agglomerates that can be easily resuspended by traffic turbulence (Meza-Figueroa et

361

al., 2016), in this work, the association of rutile NPs with PAHs could be attributed to

362

mixing sources including traffic. Calderón-Garcidueñas et al. (2019) reported iron-rich

363

combustion- and friction-derived NPs in Mexico City, highlighting the contribution of

364

non-exhaust sources; in their work, the geometric mean diameter (nm) varies from 5 to

365

50 nm with main abundances around 10-20 nm.

366

TiO2 has three different crystalline forms (rutile, anatase, and brookite). Rutile and

367

anatase have tetragonal unit cell parameters (a=b≠c), while brookite has orthorhombic

368

unit cell parameters (a≠b≠c). For rutile a=b= 4.505 to 4.593Å and c= 2.959 to 3.027 Å;

369

for anatase a=b= 3.730 to 3.828 Å and c= 9.090 to 9.514 Å (Table 2). Brookite has a

370

significantly higher value of a= 9.166 to 9.182 Å when compared to those reported for

371

rutile and anatase. The distinction among the mineralogical types is achievable because

372

rutile has a value of c= 2.959 to 3.027 Å, which is much lower than those reported for

373

anatase (c= 9.09 to 9.514 Å). High Resolution-Transmission Electron Microscopy (HR-

374

TEM) was conducted in the TiO2 crystals in this study, allowing the determination of the

375

following interplanar distance 2.5 Å (101) and 3.2 Å (110) (Figure 6A and B). Our data,

376

when compared to those provided by powder diffraction files database (PDF in Table 2,

377

Bruker Diffrac.eva plus software) showed that rutile is the crystalline form of TiO2-NPs

378

with the unit cell parameters: a= 4.508 Å and c= 3.02 Å. This result is relevant since,

379

among the three mineralogical TiO2-types, rutile NPs genotoxicity is documented as the

380

highest (Uboldi et al., 2016).

381

382

4. Discussion

383

The physical and chemical properties of rutile are well documented. The environmental

384

fate of Ti-NPs is mainly known in aquatic media (Gondikas et al., 2018) and soils (Loosli

385

et al., 2019) with scarce reports in atmospheric environments (Silva et al., 2020).

386

Most published research deals with the toxicity of TiO2 NPs in aquatic environments (Shi

387

et al., 2019; Kong et al., 2019), but these studies have been performed with commercially

388

available compounds and not from heterogeneous and complex atmospheric

389

environmental samples. Furthermore, rutile genotoxicity has been widely reported, but all

390

of these studies used purchased NPs and did not report its presence in the environmental

391

samples (Ghosh et al., 2017; Jalili et al., 2018; Rizk et al., 2017; Saquib et al., 2012;

392

Tavares et al., 2014; Uboldi et al., 2016). TiO2 is typically reported as the bulk

393

composition within atmospheric dust. As TiO2 is a major element of rock-forming

394

minerals, for example, in perovskite (CaTiO3), titanite (CaTiO2(SiO5)), and ilmenite

395

(FeTiO2), it can be expected to appear in many environmental samples. In most cases, it

396

is part of the structural formulae of minerals (amphiboles, micas, etc.) and not as rutile

397

forms. TiO2 NPs (as a mixture of rutile and anatase) from construction sites have been

398

found in sludge and effluents at a wastewater treatment plant (Oliveira et al., 2019; Tong

399

et al., 2015). To our knowledge, our study would be the first report of rutile NPs in

400

samples collected from the urban atmosphere at pedestrian levels.

401

In addition to rutile NPs, polycyclic aromatic hydrocarbons (PAHs) were detected. Their

402

presence is relevant as they associate with NPs, such as rutile, to be further incorporated

403

in the atmosphere. PAHs and PAH-derivatives sources are traffic emissions industrial

404

plants, waste incinerators, and open burning. Genotoxicity of rutile NPs could be

405

increased by the association with carcinogenic and mutagenic PAHs (Billet et al., 2007;

406

Khpalwak et al., 2019; Saint-Georges et al., 2007).

407

We also found Fe-Al-TiO2 particles as microaggregates associated with rutile (Schindler

408

and Hochella, 2016). NPs rapid aggregation can be due to the surface charge interactions,

409

forming microaggregates with organic matter (Calderón-Garcidueñas et al., 2019; Pradas

410

Del Real et al., 2018).

411

Chronic inhalation studies have shown that aggregated ultrafine particles of TiO2 are

412

more likely to induce tumor production in rats more effectively than larger particles, and

413

it has been suggested that a toxicological impact due to their physicochemical properties

414

may be an issue. Studies in animal models have shown global DNA hypo-methylation in

415

liver tissues after oral administration of TiO2 NPs and pulmonary inflammation, raising

416

concern about the safety of these materials (El Dine et al., 2018; Sun et al., 2012).

417

Particles of various TiO2 composites may have toxicities which vary depending upon

418

crystal structure, particle size, and surface characteristics (Warheit, 2013). The crystalline

419

composition of TiO2 NPs influences its toxicity in the water flea Daphnia magna when

420

based on the mass concentration (Bundschuh et al., 2018; Seitz et al., 2014). TiO2 has

421

also been associated with vascular dysfunction that may contribute to ischemic cardiac

422

events (LeBlanc et al., 2009).

423

Moreover, the size and chemical composition of PM varies greatly and depends on

424

factors such as combustion sources, climate, season, and urban or industrial pollution

425

(Karagulian et al., 2015). Communities in urban areas face complex environmental risks

426

due to development challenges such as the reduction of green spaces, increasing surface

427

runoff due to impermeable cover by pavement, and air pollution. An arid zone, such as

428

the Sonoran Desert where Hermosillo is located, has a particular airborne composition

429

affected mainly by dust storms, which trigger numerous small particles into the air, and

430

impacts its quality. Moreover, high surface temperature causes a loss in the soil humidity,

431

particle cohesion, and consequently fine and ultrafine particles are released to the

432

atmosphere (Crooks et al., 2016; Renzi et al., 2018).

433

Tackling the problem of increased PM represents a tremendous scientific as well as

434

economic challenge. Studies showed that up to 50% of PM2.5 originates from traffic

435

emissions and resuspension of road dust (Watson and Chow, 2001). The ambient

436

concentrations of airborne PM10 and PM2.5 are regulated through ambient air quality

437

standards but usually exceed the recommended values in many countries, with alarmingly

438

high concentrations in several big cities and rural areas, particularly in some developing

439

countries. Moreover, there are currently no air quality guidelines for NPs. For this

440

reason, ultrafine PM should be monitored by international organizations to control and

441

reduce sources that impact outdoor air pollution to avoid public exposures to harmful

442

contaminants and human health complications. Our protocol could be a valuable resource

443

to accomplishing the goal of a better air quality assessment.

444 445

5. Conclusion

446

The characterization of fine and ultrafine PM is of particular importance due to prolonged

447

residence time in the atmosphere and the documented adverse effects of human exposure.

448

In this study, a modification of the standard protocols for NPs extraction is proposed for

449

the identification of nanocrystals of rutile (TiO2) and PAHs.

450

Rutile crystals were detected in nanometric sizes, which increase the risk of diseases

451

caused by this hazardous material. Rutile is a commonly found form of TiO2 (compared

452

to anatase and brookite), and it is used worldwide in multiple applications. Rutile found

453

in atmospheric dust is possibly due to resuspension primarily by wind, pedestrian and

454

vehicle traffic.

455

The spatial distribution of Ti in road dust showed accumulation areas controlled by

456

erosion, traffic, and the topography of the urban area. Such places should be monitored as

457

dust emission sources. Given the public health implications of the presence of rutile NPs

458

and PAHs in atmospheric dust further studies are recommended.

459 460

6. Conflicts of interest

461

The authors declare no competing financial interests.

462 463

7. Acknowledgments

464

We gratefully acknowledge the use of TEM facilities at the TEM Laboratory of the

465

University of Sonora. We also thank Lilián F. Hernández-Valdez and Cristian Hurtado-

466

Irigoyen for their support during sample collection; Andre-i Sarabia-Sainz for his

467

assistance in the graphic abstract and Alejandro Huerta-Saquero for his comments on this

468

manuscript.

469 470

8. Funding

471

National Council for Science and Technology in Mexico (CONACYT) Grant A1-S-

472

29697 to Diana Meza-Figueroa. Ana L. Gallego-Hernández was funded by a Repatriation

473

fellowship from CONACyT.

474 475

9. References

476

Adachi, K., Buseck, P.R., 2010. Hosted and free-floating metal-bearing atmospheric

477

nanoparticles in Mexico city. Environ Sci Technol. 44, 2299–2304.

478

https://doi.org/10.1021/es902505b

479

Agarwal, A.K., et al. 2018. Mutagenicity and cytotoxicity of particulate matter emitted

480

from biodiesel-fueled engines. Environ Sci Technol. 52, 14496–14507.

481

https://doi.org/10.1021/acs.est.8b03345

482

Ali, M.U., et al. 2018. A systematic review on global pollution status of particulate

483

matter-associated potential toxic elements and health perspectives in urban

484

environment. Environ Geochem Health 0123456789.

485

https://doi.org/10.1007/s10653-018-0203-z

486 487

Alves, C.A., et al. 2018. Chemical profiling of PM10 from urban road dust. Sci Total Environ. 634, 41–51. https://doi.org/10.1016/j.scitotenv.2018.03.338

488

Amato, F., et al. 2009. Quantifying road dust resuspension in urban environment by

489

multilinear engine: A comparison with PMF2. Atmos Environ. 43, 2770–2780.

490

https://doi.org/10.1016/j.atmosenv.2009.02.039

491

Amato, F., et al. 2011. Sources and variability of inhalable road dust particles in three

492

European cities. Atmos Environ. 45, 6777–6787.

493

https://doi.org/10.1016/j.atmosenv.2011.06.003

494

Andreau, K., et al. 2012. Health and cellular impacts of air pollutants: From

495

cytoprotection to cytotoxicity. Biochem Res Int. 2012, 1–18.

496

https://doi.org/10.1155/2012/493894

497

Apeagyei, E., et al. 2011. Distribution of heavy metals in road dust along an urban-rural

498

gradient in Massachusetts. Atmos Environ. 45, 2310–2323.

499

https://doi.org/10.1016/j.atmosenv.2010.11.015

500

Billet, S., et al. Aboukais, A., Shirali, P., 2007. Ambient particulate matter (PM2.5):

501

Physicochemical characterization and metabolic activation of the organic fraction in

502

human lung epithelial cells (A549). Environ Res. 105, 212–223.

503

https://doi.org/10.1016/j.envres.2007.03.001

504 505 506

Bowen, H.J.M. 1979. Environmental Chemistry of the Elements. Academic Press, London. Buat-Menerd, P., Chesslet, R. 1979. Variable influence of the atmospheric flux on the

507

trace element chemistry of oceanic suspended matter. Earth Planet Sci Lett. 42, 399-

508

411. https://doi.org/10.1016/0012-821X(79)90049-9

509

Bundschuh, M., et al. 2018. Nanoparticles in the environment: where do we come from,

510

where do we go to? Environ Sci Eur. 30, 2–17. https://doi.org/10.1186/s12302-018-

511

0132-6

512

Calderón-Garcidueñas, L., et al. 2019. Combustion- and friction-derived magnetic air

513

pollution nanoparticles in human hearts. Environ Res. 176, 108567.

514

https://doi.org/10.1016/j.envres.2019.108567

515 516

Chehregani, A., et al. 2004. Increasing allergy potency of Zinnia pollen grains in polluted areas. Ecotoxicol Environ Saf. 58, 267–272.

517 518

https://doi.org/10.1016/j.ecoenv.2003.12.004 Chen, J., et al. 2014. Characterization of polycyclic aromatic hydrocarbons using Raman

519

and surface enhanced Raman Spectroscopy. J Raman Spectrosc. 46, 64–69.

520

https://doi.org/10.1002/jrs.4612

521 522 523 524 525

COESPO, 2015. Hermosillo: Indicadores demográficos y socioeconómicos. Encuesta Intercensal 2015. INEGI. http://www.coespo.sonora.gob.mx/ Cornu, S., et al. 1999. Evidence of titanium mobility in soil profiles, Manus, Central Amazonia. Geoderma. 91, 281-295. https://doi.org/10.1016/S0016-7061(99)00007-5 Crooks, J.L., et al. 2016. The association between dust storms and daily non-accidental

526

mortality in the United States, 1993-2005. Environ Health Perspect. 124, 1735–

527

1743. https://doi.org/10.1289/EHP216

528

Cullen, E., et al. 2010. Simulation of the subsurface mobility of carbon nanoparticles at

529

the field scale. Adv Water Res. 33, 361-371.

530

https://doi.org/10.1016/j.advwatres.2009.12.001

531

Dias, M.P., et al. 2013. Morphological characterization of soil clay fraction in nanometric

532

scale. Powder Technol. 241, 36-42. https://doi.org/10.1016/j.powtec.2013.03.006

533

El Dine, F.M.M.B., et al. 2018. Evaluation of epigenetic changes of liver tissue induced

534

by oral administration of Titanium dioxide nanoparticles and possible protective role

535

of Nigella Sativa oil, in adult male albino rats. Nanomed J. 5, 192–198.

536

https://doi.org/10.22038/nmj.2018.05.00002

537

Evans, K.A., et al. 2014. Increased ultrafine particles and carbon monoxide

538

concentrations are associated with asthma exacerbation among urban children.

539

Environ Res. 129, 11–19. https://doi.org/10.1016/j.envres.2013.12.001

540

Farjana, S.H., et al. 2018. Towards sustainable TiO2 production: An investigation of

541

environmental impacts of ilmenite and rutile processing routes in Australia. J Clean

542

Prod. 196, 1016–1025. https://doi.org/10.1016/j.jclepro.2018.06.156

543

Franchini, M., Mannucci, P.M., 2007. Short-term effects of air pollution on

544

cardiovascular diseases: Outcomes and mechanisms. J Thromb Haemost. 5, 2169–

545

2174. https://doi.org/10.1111/j.1538-7836.2007.02750.x

546

Frank, O., et al. 2012. Raman spectra of titanium dioxide (anatase, rutile) with identified

547

oxygen isotopes (16, 17, 18). Phys Chem Chem Phys. 14, 14567–14572.

548

https://doi.org/10.1039/c2cp42763j

549

Gázquez, M.J., et al. 2014. A review of the production cycle of titanium dioxide pigment.

550

Mater Sci Appl. 5, 441–458.

551

https://doi.org/http://dx.doi.org/10.4236/msa.2014.57048

552

Gondikas, A., et al. 2018. Where is the nano? Analytical approaches for the detection and

553

quantification of TiO2 engineered nanoparticles in surface waters. Environ Sci Nano.

554

5, 313. https://doi.org/10.1039/C7EN00952F

555

Gosh, M., et al. 2017. Cyto-genotoxic and DNA methylation changes induced by

556

different crystal phases of TiO2-np in bronchial epithelial (16-HBE) cells. Mutat

557

Res. 796, 1-12. https://doi:10.1016/j.mrfmmm.2017.01.003

558

Hakanson, L., 1980. An ecological risk index for aquatic pollution control. A

559

sedimentological approach. Water Res. 14, 975-1001. https://doi.org/10.1016/0043-

560

1354(80)90143-8

561 562

Hartland, A. et al. 2013. The environmental significance of natural nanoparticles. Nat Edu K. 2013, 4, no. 8, 1-7.

563

Jakubiak-Lasocka, J., et al. 2015. The influence of particulate matter on respiratory

564

morbidity and mortality in children and infants. Adv Exp Med Biol 10, 39–48.

565

https://doi.org/10.1007/5584

566

Jalili, P., et al. 2018. Investigation of the in vitro genotoxicity of two rutile TiO2

567

nanomaterials in human intestinal and hepatic cells and evaluation of their

568

interference with toxicity assays. NanoImpact. 11, 69-81.

569

https://doi.org/10.1016/j.impact.2018.02.004

570

Jiménez, E., et al. 2010. Role of Saharan dust in the relationship between particulate

571

matter and short-term daily mortality among the elderly in Madrid (Spain). Sci Total

572

Environ. 408, 5729–5736. https://doi.org/10.1016/j.scitotenv.2010.08.049

573

Jin, X., et al. 2018. Mitochondrial damage mediated by ROS incurs bronchial epithelial

574

cell apoptosis upon ambient PM2.5 exposure. J Toxicol Sci. 43, 101–111.

575

https://doi.org/10.2131/jts.43.101

576

Karagulian, F., et al. 2015. Contributions to cities’ ambient particulate matter (PM): A

577

systematic review of local source contributions at global level. Atmos Environ. 120,

578

475–483. https://doi.org/10.1016/j.atmosenv.2015.08.087

579

Khpalwak, W., J et al. 2019. Polycyclic aromatic hydrocarbons in urban road dust,

580

Afghanistan: Implications for human health. Chemosphere. 218, 517–526.

581

https://doi.org/10.1016/j.chemosphere.2018.11.087

582

Kim, K.H., et al. 2015. A review on the human health impact of airborne particulate

583

matter. Environ Int. 74, 136–143. https://doi.org/10.1016/j.envint.2014.10.005

584

Kong, H., et al. 2019. Nano-TiO2 impairs digestive enzyme activities of marine mussels

585

under ocean acidification. Chemosphere. 237, 124561.

586 587 588 589

https://doi.org/10.1016/j.chemosphere.2019.124561 Krug, H.F., Wick, P., 2011. Nanotoxicology: An interdisciplinary challenge. Angew Chemie Int Ed Engl. 50, 1260–1278. https://doi.org/10.1002/anie.201001037 Kusin, F.M., et al. 2019. Geo-ecological evaluation of mineral, major and trace elemental

590

composition in waste rocks, soils and sediments of a gold mining area and potential

591

associated risks. Catena 183, 104229. https://doi.org/10.1016/j.catena.2019.104229

592

LeBlanc, A., et al. 2009. Nanoparticle inhalation impairs endothelium-dependent

593

vasodilation in subepicardial arterioles. J Toxicol Env Heal A. 72, 1576–1584.

594

https://doi.org/10.1080/15287390903232467.

595

Lee, J., et al. 2010. Nanomaterials in the construction industry: A review of their

596

applications and environmental heatlh and safety considerations. ACS Nano. 4,

597

3580–3590. https://doi:10.1021/nn100866w

598

Longhin, E., et al. 2013. Season linked responses to fine and quasi-ultrafine Milan PM in

599

cultured cells. Toxicol In Vitro. 27, 551–559.

600

https://doi.org/10.1016/j.tiv.2012.10.018

601

Longhin, E., et al. 2016. Integrative transcriptomic and protein analysis of human

602

bronchial BEAS-2B exposed to seasonal urban particulate matter. Environ Pollut.

603

209, 87–98. https://doi.org/10.1016/j.envpol.2015.11.013

604

Longhin, E., et al. 2018. Milan winter fine particulate matter (wPM2.5) induces IL-6 and

605

IL-8 synthesis in human bronchial BEAS-2B cells, but specifically impairs IL-8

606

release. Toxicol In Vitro. 52, 365–373. https://doi.org/10.1016/j.tiv.2018.07.016

607

Loosli, F., et al. 2019. Improved extraction efficiency of natural nanomaterials in soils to

608

facilitate their characterization using a multimethod approach. Sci total Environ.

609 610

677, 34-46. https://doi.org/10.1016/j.scitotenv.2019.04.301 Lu, S., et al. 2014. Characterization of protein expression of Platanus pollen following

611

exposure to gaseous pollutants and vehicle exhaust particles. Aerobiologia

612

(Bologna). 30, 281–291. https://doi.org/10.1007/s10453-014-9327-5

613 614 615

Meinhold, G., 2010. Rutile and its applications in earth sciences. Earth-Science Rev. 102, 1–28. https://doi.org/10.1016/j.earscirev.2010.06.001 Meza-Figueroa, D., et al. 2016. Traffic signatures in suspended dust at pedestrian levels

616

in semiarid zones: Implications for human exposure. Atmos Environ. 138, 4–14.

617

https://doi.org/10.1016/j.atmosenv.2016.05.005

618

Meza-Figueroa, D., et al. 2018. Source apportionment and environmental fate of lead

619

chromates in atmospheric dust in arid environments. Sci Total Environ. 630, 1596–

620

1607. https://doi.org/10.1016/j.scitotenv.2018.02.285

621

Miller, R.L., et al. 2006. Mineral dust aerosols in the NASA Goddard Institute for Space

622

Sciences ModelE atmospheric general circulation model. J Geophys Res. 111, 1–19.

623

https://doi.org/10.1029/2005JD005796

624

Moreno-Rodríguez, V., et al. 2015. Historical trends and sources of TSP in a Sonoran

625

desert city: Can the North America Monsoon enhance dust emissions? Atmos

626

Environ. 110, 111–121. https://doi.org/10.1016/j.atmosenv.2015.03.049

627

Morillas, H., et al. 2018a. In-situ analytical study of bricks exposed to marine

628

environment using hand-held X-ray fluorescence spectrometry and related

629

laboratory techniques. Spectrochim Acta Part B At Spectrosc. 146, 28-35.

630

https://doi.org/10.1016/j.sab.2018.04.020

631

Morillas, H., et al. 2018b. Composition and porosity study of original and restoration

632

materials included in a coastal historical construction. Constr Build Mater. 178, 384-

633

392. https://doi.org/10.1016/j.conbuildmat.2018.05.168

634 635 636

Muller G. 1969. Index of geoaccumulation in sediments of the Rhine River. Geojournal 2, 108-118. Nagar, J.K., et al. 2014. A review on airborne particulate matter and its sources, chemical

637

composition and impact on human respiratory system. 5, 447–463

638

https://doi.org/10.6088/ijes.2014050100039

639

Navarro E., et al. 2008. Environmental behavior and ecotoxicity of engineered

640

nanoparticles to algae, plants, and fungi. Ecotoxicology 17: 372-386.

641

https://doi.org/10.1007/s10646-008-0214-0

642

Nikolla P. Qafoku, 2008. Terrestrial nanoparticles and their controls on soil-/geo-

643

processes and reactions. Adv Agron. 107, 33-91. https://doi.org/10.1016/S0065-

644

2113(10)07002-1

645 646 647

Oberdörster, G., 2001. Pulmonary effects of inhaled ultrafine particles. Int Arch Occup Environ. Health. 74, 1–8. https://doi.org/10.1007/s004200000185 Oberdörster, G., et al. 2005. Nanotoxicology: an emerging discipline evolving from

648

studies of ultrafine particles. Environ Health Perspect. 113, 823–839.

649

https://doi.org/10.1289/ehp.7339

650

Oliveira, M.L.S., et al. 2019. Nanoparticles from construction wastes: A problem to

651

health and the environment. J Clean Prod. 219, 236–243.

652

https://doi.org/10.1016/j.jclepro.2019.02.096

653

Perrone, M.G., et al. 2010. Seasonal variations in chemical composition and in vitro

654

biological effects of fine PM from Milan. Chemosphere. 78, 1368–1377.

655 656

https://doi.org/10.1016/j.chemosphere.2009.12.071 Piccinno, F., et al. 2012. Industrial production quantities and uses of ten engineered

657

nanomaterials in Europe and the world. J Nanopart Res. 144, 1109.

658

https://doi.org/10.1007/s11051-012-1109-9

659

Pope, C.A., et al. 2004. Cardiovascular mortality and long-term exposure to particulate

660

air pollution: Epidemiological evidence of general pathophysiological pathways of

661

disease. Circulation. 109, 71–77.

662

https://doi.org/10.1161/01.CIR.0000108927.80044.7F

663

Pradas Del Real, A.E., et al. 2018. Searching for relevant criteria to distinguish natural

664

vs. anthropogenic TiO2 nanoparticles in soils. Environ Sci Nano. 5, 2853–2863.

665

https://doi.org/10.1039/c8en00386f

666

Ravindra, K., et al. 2008. Atmospheric polycyclic aromatic hydrocarbons: Source

667

attribution, emission factors and regulation. Atmos Environ. 42, 2895–2921.

668

https://doi.org/10.1016/j.atmosenv.2007.12.010

669

Renzi, M., F et al. 2018. Short-term effects of desert and non-desert PM10 on mortality in

670

Sicily, Italy. Environ Int. 120, 472–479. https://doi.org/10.1016/j.envint.2018.08.016

671 672 673

Ribeiro, H., et al. 2015. Characterisation of particulate matter on airborne pollen grains. Environ Pollut. 206, 7–16. https://doi.org/10.1016/j.envpol.2015.06.015 Ribeiro, J., et al. 2010. Identification of nanominerals and nanoparticles in burning coal

674

waste piles from Portugal. Sci Total Environ. 408, 6032-6041.

675

https://doi.org/10.1016/j.scitotenv.2010.08.046

676 677

Rizk, M.Z., et al., 2017. Toxicity of titanium dioxide nanoparticles: Effect of dose and time on biochemical disturbance, oxidative stress and genotoxicity in mice. Biomed

678

Pharmacother. 90, 466-472. http://dx.doi.org/10.1016/j.biopha.2017.03.089

679

Robichaud, C.O., et al. 2009. Estimates of upper bounds and trends in nano-TiO2

680

production as a basis for exposure assessment. Environ Sci Technol. 43, 4227–4233.

681

https://doi.org/10.1021/es8032549

682

Rönkkö, T., et al. 2017. Traffic is a major source of atmospheric nanocluster aerosol.

683

Proc Natl Acad. Sci. U. S. A. 114, 7549–7554.

684

https://doi.org/10.1073/pnas.1700830114

685

Saint-Georges, F., et al. 2007. Gene expression induction of volatile organic compound

686

and/or polycyclic aromatic hydrocarbon-metabolizing enzymes in isolated human

687

alveolar macrophages in response to airborne particulate matter (PM2.5). Toxicology.

688

105, 220–230. https://doi.org/10.1016/j.tox.2007.11.016

689

Saquib, Q., et al. 2012. Titanium dioxide nanoparticles induced cytotoxicity, oxidative

690

stress and DNA damage in human amnion epithelial (WISH) cells. Toxicol In Vitro.

691

26, 351–361. https://doi.org/10.1016/j.tiv.2011.12.011

692

Schindler, M., and Hochella, M.F. 2016. Nanomineralogy as a new dimension in

693

understanding elusive geochemical processes in soils: The case of low-solubility-

694

index elements. Geology. 44: 515-518. https://doi.org/10.1130/G37774.1

695

Sebesta, M., et al. 2020. Partitioning and stability of ionic, nano- and microsized zinc in

696

natural soil suspensions. Sci Total Environ 700, 134445.

697

https://doi.org/10.1016/j.scitotenv.2019.134445

698 699 700

Sedghy, F., et al. 2018. Interaction between air pollutants and pollen grains: The role on the rising trend in allergy. Rep Biochem Mol Biol. 6, 219–224. Seitz, F., et al. 2014. Size-, surface- and crystalline structure composition-related effects

701

of titanium dioxide nanoparticles during their aquatic life cycle. Sci Total Environ.

702

493, 891–897. https://doi.org/10.1016/j.scitotenv.2014.06.092

703 704

Sharma, R.K., et al. 2013. Chemical characterization of airborne particulates of Kathmandu, Nepal. Res J Chem. Sci. 3, 88–96.

705

Shi, W., et al., 2019. Ocean acidification increases the accumulation of titanium dioxide

706

nanoparticles (nTiO2) in edible bivalve mollusks and poses a potential threat to

707

seafood safety. Sci Rep. 9: 3516. https://doi.org/10.1038/s41598-019-40047-1

708

Silva F.O., et al. 2020. Multiple hazardous elements in nanoparticulate matter from a

709

Caribbean industrialized atmosphere. Chemosphere 239, 124776.

710

https://doi.org/10.1016/j.chemosphere.2019.124776

711 712 713

Slezakova, K., et al. 2013. Atmospheric nanoparticles and their impacts on public health. Current Topics in Public Health. https://doi.org/10.5772/54775 Sun, Q., et al. 2012. Pulmotoxicological effects caused by long-term titanium dioxide

714

nanoparticles exposure in mice. J Hazard Mater. 236, 47–53.

715

https://doi:10.1016/j.jhazmat.2012.05.072.

716

Sutherland, R.A., 2000. Bed sediment-associated trace metals in an urban stream, Oahu,

717

Hawaii Environ Geol. 39, 611-627. https://doi.org/10.1007/s002540050473

718

Taboada T., et al. 2006. Particle-size fractionation of titanium and zirconium during

719

weathering and pedogenesis of granitic rocks in NW Spain. Geoderma 131, 218-

720

236. https://doi.org/10.1016/j.geoderma.2005.03.025

721

Tanner, P.A., et al. 2008. Fingerprinting metals in urban street dust of Beijing, Shanghai,

722

and Hong Kong. Environ Sci Technol. 42, 7111–7117.

723

https://doi.org/10.1021/es8007613

724

Tavares, A.M., et al. 2014. Genotoxicity evaluation of nanosized titanium dioxide,

725

synthetic amorphous silica and multi-walled carbon nanotubes in human

726

lymphocytes. Toxicol In Vitro. 28, 60-69. https://doi:10.1016/j.tiv.2013.06.009

727

Taylor, S. R., and McLennan, S. M. 1985. The continental crust: its composition and

728 729

evolution. Tong, T., et al. 2015. Spectroscopic characterization of TiO2 polymorphs in wastewater

730

treatment and sediment samples. Environ Sci Technol Lett. 2, 12–18.

731

https://doi.org/10.1021/ez5004023

732

Uboldi, C., et al. 2016. Role of the crystalline form of titanium dioxide nanoparticles:

733

rutile, and not anatase, induces toxic effects in Balb/3T3 mouse fibroblasts. Toxicol

734

In Vitro. 31, 137-145. https://doi:10.1016/j.tiv.2015.11.005

735 736

USEPA, 2007. Field portable X-ray fluorescence spectrometry for the determination of elemental concentrations in soil and sediment. 1-32.

737

Valotto, G., et al. 2015. Environmental and traffic-related parameters affecting road dust

738

composition: A multi-technique approach applied to Venice area (Italy). Atmos

739

Environ. 122, 596–608. https://doi.org/10.1016/j.atmosenv.2015.10.006

740

Von Der Kammer, F., et al. 2012. Analysis of engineered nanomaterials in complex

741

matrices (environment and biota): General considerations and conceptual case

742

studies. Nanomater Environ. 31, 32–49. https://doi.org/10.1002/etc.723

743

Wagner, S., et al. 2014. Spot the Difference: Engineered and Natural Nanoparticles in

744

the Environment-Release, Behavior, and Fate. Angewandte Chemie International

745

Edition. 31. DOI: 10.1002/anie.201405050

746

Warheit, D.B., 2013. How to measure hazards/risks following exposures to nanoscale or

747

pigment-grade titanium dioxide particles. Toxicol Lett. 220, 193–204.

748

https://doi.org/10.1016/j.toxlet.2013.04.002

749

Watson, J.G., Chow, J.C., 2001. Source characterization of major emission sources in the

750

Imperial and Mexicali Valleys along the US/Mexico border. Sci Total Environ. 276,

751

33–47. https://doi.org/10.1016/S0048-9697(01)00770-7

752

Weir, A., et al. 2012. Titanium doxide nanoparticles in food and personal care products.

753

Environ Sci Technol. 46, 2242–2250. https://doi.org/10.1021/es204168d

754

WHO Regional Office for Europe, 2013. Health effects of particulate matter. Policy

755 756 757 758

implications for countries in eastern Europe, Caucasus and central Asia. 1–13. Wilcox, J.D., 1956. Isokinetic flow and sampling. J Air Pollut Control Assoc. 5, 226– 245. https://doi.org/10.1080/00966665.1956.10467715 Wu S., et al. 2020. Identification and quantification of titanium nanoparticles in surface

759

water: A case study in Lake Taihu, China. J Hazard Mater. 382, 121045.

760

https://doi.org/10.1016/j.jhazmat.2019.121045

761

Zhang, Y., et al. 2019. Fine particle-bound polycyclic aromatic hydrocarbons (PAHs) at

762

an urban site of Wuhan, central China: Characteristics, potential sources and cancer

763

risks apportionment. Environ Pollut. 246, 319-327.

764

https://doi.org/10.1016/j.envpol.2018.11.111

765

Zou, Y., et al. 2017. Physicochemical properties, in vitro cytotoxic and genotoxic effects

766

of PM1.0 and PM2.5 from Shanghai, China. Environ Sci Pollut Res. 24, 19508–

767

19516. https://doi.org/10.1007/s11356-017-9626-9

768 769

770

Figure Captions

771

Figure 1.

772

Spatial distribution of titanium (Ti) in Hermosillo, Mexico. A) Road dust sample

773

locations (n=226) and local geochemical background sites (LGB) (n=10) are indicated by

774

black and white triangles, respectively. Titanium (Ti) concentrations are shown in mg.kg-

775

1

776

from road dust samples. LGB: Local geochemical background corresponds to

777

uncontaminated soils collected outside the urbanized area. Ti concentration of LGB is

778

indicated by the red dashed line. Anthropogenic sources refer to particles from a non-

779

natural origin.

. A star indicates atmospheric sample site. B) Histogram of Ti concentrations obtained

780 781

Figure 2.

782

Titanium (Ti) and Iron (Fe) aggregates from atmospheric samples. SEM-EDS analysis

783

from airborne particles on the filter surface.

784 785

Figure 3.

786

Size distribution analysis and chemical characterization of PM1.0. Transmission Electron

787

Microscopy (TEM) images of airborne PM. A) Total particles. B) NPs aggregates. C)

788

Particles size distribution analysis, n=628. D) EDS energy peaks. El: element; %Wt:

789

weight percentage.

790 791

Figure 4.

792

Chemical identification of titanium (Ti) and oxygen (O) in airborne PM1.0. A) Scanning

793

Electron Microscopy (SEM) image; red circle indicates the site for EDS analysis B) EDS

794

spectrum of Ti and O. El: element; %Wt: weight percentage.

795 796

Figure 5.

797

Identification of titanium dioxide (TiO2) by Confocal Raman Spectroscopy analysis. A)

798

Optical image. B) Raman spectroscopy image. C) Raman spectra clustered 1-4. D)

799

Localization of the clustered classification.

800 801

Figure 6.

802

Rutile crystals NPs found on airborne PM. A) High Resolution-Transmission Electron

803

Microscopy (HR-TEM) image of rutile showing 0.25 nm spaced lattice fringes, which

804

correspond to (110) planes of rutile. B) FFT Diffractogram of A). The zone axis and the

805

rutile spots 101 and 110 are indicated.

Table 1. The titanium content in road dust (RD); n: number of samples analyzed in each study. LGB: Local geochemical background corresponds to uncontaminated soils collected outside the urbanized area. Locality

n

Maximum (mg.kg-1)

Minimum (mg.kg-1)

Hermosillo, Mexico

226

8409

1248

Study point

1

4211

LGB

10

Massachusetts, USA

85

Venice, Italy

16

Zurich, Switzerland

8

Barcelona, Spain

9

Girona, Spain

Mean SD

± Reference

3447±1088

Present study

4187

4201±12

Present study

3890

2992

3390±350

Present study

4941

742

2029±1016

(Apeagyei et al., 2011)

3700

2770

3210

(Valotto et al., 2015)

-

-

1488±561

(Amato et al., 2011)

-

-

2964±1207

(Amato et al., 2011)

6

-

-

2113±395

(Amato et al., 2011)

Beijing, China

8

-

-

4400±600

(Tanner et al., 2008)

Shanghai, China

9

-

-

3200±500

(Tanner et al., 2008)

Hong Kong

8

-

-

2300±600

(Tanner et al., 2008)

Table 2. Rutile, anatase, and brookite unit cell parameters obtained by HR-TEM, and comparison with values reported by XRD Powder Diffraction Files database. Syn: synthetic. Source

ID

Quality

Status

Name

Mineral Name

Formula

Crystal System

a (Å)

c (Å)

Minerals, Metals , Alloy & Corrotion

PDF 820514

Calculated

Primary

Titanium Dioxide

Rutile

TiO2

Tetragonal

4.508

3.027

Metals , Alloy & Corrotion

PDF 340180

Calculated

Excluded

Titanium Dioxide

Rutile, Syn.

TiO2

Tetragonal

4.593

2.959

Minerals & Metals

PDF 010562

Blank

Excluded

Titanium Dioxide

Anatase

TiO2

Tetragonal

3.730

9.370

Mineral, Metal, Alloy, Pigment, Corrosion

PDF 211272

Star (*)

Primary

Titanium Dioxide

Anatase, Syn. TiO2

Tetragonal

3.785

9.514

Minerals, Metals , Alloy & Corrotion

PDF 861155

Calculated

Primary

Titanium Dioxide

Anatase, Syn. TiO2

Tetragonal

3.807

9.090

Metals , Alloy & Corrotion

PDF 020514

Low Precision

Excluded

Titanium Dioxide

Brookite

TiO2

Orthorhombic

9.166

5.135

Mineral, Metal & Alloy

PDF 150875

Star (*)

Excluded

Titanium Dioxide

Brookite

TiO2

Orthorhombic

9.182

5.143

Mineral, Metal & Alloy

PDF 030380

Blank

Excluded

Titanium Dioxide

Brookite

TiO2

Orthorhombic

9.166

5.135

TiO2

HR-TEM results (present study) Tetragonal 4.508

3.02

Ti (mg.kg-1)

6000 6500 7000 7500 8000 8500

5500

LGB

4000 4500 5000

2500 3000 3500

1000 1500 2000

A

500

Frequency

Figure 1

B

50 Anthropogenic

40

30

20

10 * ** *

0

Figure 2

Fe-Ti

Figure 3 140

A C

120 n = 636 Mean = 77.5 ± 66.3

Frequency

100 80 60 40 20 0 B

D

Diameter (nm)

350 300

El

% Wt

200

C

90.41

150

O

8.36

Si

0.71

Ti

0.52

250

100 50 0 1

2

3

4

5 keV

6

7

8

9

Figure 4 A

B C

El

% Wt

Ti

76.7

O

23.3

O

Ti Ti

Ti

5µm

0

1

2

3

4

5

6

7

8

9

10 11 12 13 14

Figure 5 A

B

C

D Cluster 1 Cluster 2 Cluster 3 Cluster 4

500 500

1000 1500 2000 2500 1000 1500 2000 2500 -1 Raman Shift (cm ) Raman Shift (cm-1)

3000 3000

Figure 6 A

B

(110)

(101)

2.5 Å (101) [-111]

Gallego-Hernández et al “Identification of inhalable rutile and polycyclic aromatic hydrocarbons (PAHs) nanoparticles in the atmospheric dust.” Highlights •

Synthetic rutile as NPs aggregates were identified in environmental samples

•

Rutile NPs linked to PAHs were found in the airborne PM at pedestrian levels

•

Ultrafine fraction of airborne dust showed particle size within the 41-60 nm range

Author Contribution Statement

Ana L. Gallego-Hernández: Methodology, Validation, Writing - Original Draft. Diana MezaFigueroa: Conceptualization, Methodology, Investigation, Writing - Original Draft, Supervision, Funding acquisition, Writing - Original Draft. Judith Tanori: Data Curation. Mónica AcostaElías: Methodology, Data Curation. Belem González-Grijalva: Methodology, Data Curation. Juan F. Maldonado-Escalante: Methodology. Sarai Rochín-Wong: Methodology. Diego Soto-Puebla: Writing - Review & Editing. Sofia Navarro-Espinoza: Methodology, Visualization, Review & Editing. Roberto Ochoa-Contreras: Methodology, Data Curation. Martín Pedroza-Montero: Conceptualization, Formal analysis, Investigation, Writing Original Draft, Supervision, Funding acquisition, Writing - Original Draft.

Declaration of interests ☐ The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

☐ The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: