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What contributes to human body burdensof phthalate esters?An experimental approachGeorgios Giovanoulis
Academic dissertation for the Degree of Doctor of Philosophy in Applied EnvironmentalScience at Stockholm University to be publicly defended on Friday 1 September 2017 at 10.00in Nordenskiöldsalen, Geovetenskapens hus, Svante Arrhenius väg 12.
AbstractPhthalate esters (PEs) and alternative plasticizers used as additives in numerous consumer products are continuouslyreleased into the environment leading to subsequent human exposure. The ubiquitous presence and potential adverse healtheffects (e.g. endocrine disruption and reproductive toxicity) of some PEs are responsible for their bans or restrictions. Thishas led to increasing use of alternative plasticizers, especially cyclohexane-1,2-dicarboxylic acid diisononyl ester (DINCH).Human exposure data on alternative plasticizers are lacking and clear evidence for human exposure has previously onlybeen found for di(2-ethylhexyl) terephthalate (DEHTP) and DINCH, with increasing trends in body burdens. In this thesis,a study population of 61 adults (age: 20–66; gender: 16 males and 45 females) living in the Oslo area (Norway) was studiedfor their exposure to plasticizers. Information on sociodemographic and lifestyle characteristics that potentially affect theconcentrations of PE and DINCH metabolites in adults was collected by questionnaires. Using the human biomonitoringapproach, we evaluated the internal exposure to PEs and DINCH by measuring concentrations of their metabolites in urine(where metabolism and excretion are well understood) and using these data to back-calculate daily intakes. Metabolitelevels in finger nails were also determined. Since reference standards of human metabolites for other important alternativeplasticizers apart from DINCH (e.g. DEHTP, di(2-propylheptyl) phthalate (DPHP), di(2-ethylhexyl) adipate (DEHA) andacetyl tributyl citrate (ATBC)) are not commercially available, we further investigated the urine and finger nail samplesby Q Exactive Orbitrap LC-MS to identify specific metabolites, which can be used as appropriate biomarkers of humanexposure. Many metabolites of alternative plasticizers that were present in in vitro extracts were further identified in vivoin urine and finger nail samples. Hence, we concluded that in vitro assays can reliably mimic the in vivo processes. Also,finger nails may be a useful non-invasive matrix for human biomonitoring of specific organic contaminants, but furthervalidation is needed. Concentrations of PEs and DINCH were also measured in duplicate diet, air, dust and hand wipes.External exposure, estimated based on dietary intake, air inhalation, dust ingestion and dermal uptake, was higher or equalto the back-calculated internal intake. By comparing these, we were able to explain the relative importance of differentexposure pathways for the Norwegian study population. Dietary intake was the predominant exposure route for all analyzedsubstances. Inhalation was important only for lower molecular weight PEs, while dust ingestion was important for highermolecular weight PEs and DINCH. Dermal uptake based on hand wipes was much lower than the total dermal uptakecalculated via air, dust and personal care products, but still several research gaps remain for this exposure pathway. Basedon calculated intakes, the exposure risk for the Norwegian participants to the PEs and DINCH did not exceed the establishedtolerable daily intake and reference doses, and the cumulative risk assessment for combined exposure to plasticizers withsimilar toxic endpoints indicated no health concerns for the selected population. Nevertheless, exposure to alternativeplasticizers, such as DPHP and DINCH, is expected to increase in the future and continuous monitoring is required. Findingsthrough uni- and multivariate analysis suggested that age, smoking, use of personal care products and many other everydayhabits, such as washing hands or eating food from plastic packages are possible contributors to plasticizer exposure.
Keywords: Phthalates, Alternative plasticizers, DINCH, In vivo screening, Urine, Nails, Air, Dust, Hand wipes,Duplicate diet, Predictors of exposure.
Stockholm 2017http://urn.kb.se/resolve?urn=urn:nbn:se:su:diva-143147
ISBN 978-91-7649-874-3ISBN 978-91-7649-875-0
Department of Environmental Science and AnalyticalChemistry
Stockholm University, 106 91 Stockholm
Abstract
Phthalate esters (PEs) and alternative plasticizers used as additives in
numerous consumer products are continuously released into the environment
leading to subsequent human exposure. The ubiquitous presence and potential
adverse health effects (e.g. endocrine disruption and reproductive toxicity) of
some PEs are responsible for their bans or restrictions. This has led to increasing
use of alternative plasticizers, especially cyclohexane-1,2-dicarboxylic acid
diisononyl ester (DINCH). Human exposure data on alternative plasticizers are
lacking and clear evidence for human exposure has previously only been found
for di(2-ethylhexyl) terephthalate (DEHTP) and DINCH, with increasing trends
in body burdens.
In this thesis, a study population of 61 adults (age: 20–66; gender: 16 males
and 45 females) living in the Oslo area (Norway) was studied for their exposure
to plasticizers. Information on sociodemographic and lifestyle characteristics
that potentially affect the concentrations of PE and DINCH metabolites in adults
was collected by questionnaires. Using the human biomonitoring approach, we
evaluated the internal exposure to PEs and DINCH by measuring concentrations
of their metabolites in urine (where metabolism and excretion are well
understood) and using these data to back-calculate daily intakes. Metabolite
levels in finger nails were also determined. Since reference standards of human
metabolites for other important alternative plasticizers apart from DINCH (e.g.
DEHTP, di(2-propylheptyl) phthalate (DPHP), di(2-ethylhexyl) adipate (DEHA)
and acetyl tributyl citrate (ATBC)) are not commercially available, we further
investigated the urine and finger nail samples by Q Exactive Orbitrap LC-MS to
identify specific metabolites, which can be used as appropriate biomarkers of
human exposure.
Many metabolites of alternative plasticizers that were present in in vitro
extracts were further identified in vivo in urine and finger nail samples. Hence,
we concluded that in vitro assays can reliably mimic the in vivo processes. Also,
finger nails may be a useful non-invasive matrix for human biomonitoring of
specific organic contaminants, but further validation is needed.
Concentrations of PEs and DINCH were also measured in duplicate diet, air,
dust and hand wipes. External exposure, estimated based on dietary intake, air
inhalation, dust ingestion and dermal uptake, was higher or equal to the back-
calculated internal intake. By comparing these, we were able to explain the
relative importance of different exposure pathways for the Norwegian study
population. Dietary intake was the predominant exposure route for all analyzed
substances. Inhalation was important only for lower molecular weight PEs,
while dust ingestion was important for higher molecular weight PEs and
DINCH. Dermal uptake based on hand wipes was much lower than the total
dermal uptake calculated via air, dust and personal care products, but still
several research gaps remain for this exposure pathway.
Based on calculated intakes, the exposure risk for the Norwegian participants
to the PEs and DINCH did not exceed the established tolerable daily intake and
reference doses, and the cumulative risk assessment for combined exposure to
plasticizers with similar toxic endpoints indicated no health concerns for the
selected population. Nevertheless, exposure to alternative plasticizers, such as
DPHP and DINCH, is expected to increase in the future and continuous
monitoring is required.
Findings through uni- and multivariate analysis suggested that age, smoking,
use of personal care products and many other everyday habits, such as washing
hands or eating food from plastic packages are possible contributors to
plasticizer exposure.
Keywords: Phthalates, Alternative plasticizers, DINCH, In vivo screening,
Urine, Nails, Air, Dust, Hand wipes, Duplicate diet, Predictors of exposure
Svensk sammanfattning
Ftalater (FEs) samt andra alternativa mjukgörare för plaster, som används
som tillsatser i konsumentprodukter, sprids till miljön och kan ge upphov till
ökad mänsklig exponering för dessa kemikalier. Vetskapen om närvaron och
kunskaper om potentiella hälsorisker (såsom hormon- och
reproduktionsstörningar) av vissa PE är orsaken till att de ålagts med förbud
eller restriktioner. Dessa begränsningar i användningen har gett en ökning i
bruket av alternativa mjukgörare, speciellt med avseende på cyklohexan-1,2-
dikarboxylsyra diisononylester (DINCH). Kunskaper kring mänsklig exponering
för alternativa mjukgörare saknas i stor utsträckning idag. Endast di(2-
etylhexyl), tereftalat (DEHTP) och DINCH har påvisats i tidigare studier, med
trender som visar på en ökad exponering.
I denna avhandling har en population på 61 vuxna (ålder: 20-66; kön: 16
män och 45 kvinnor) från Oslo området (Norge) studerats med avseende på
deras individuella exponering för mjukgörare. Information kring
sociodemografiska faktorer och livsstil, som kan ha betydelse för förekomsten
av PE och DINCH hos vuxna människor, samlades in i studien genom
frågeformulär. Med hjälp av human biomonitorering utvärderades förekomsten
av PE och DINCH genom att bestämma koncentrationerna av dessa metaboliter
i urinprov. Då kunskapen kring ämnesomsättning och utsöndring av dessa
substanser oftast är väl kända kunde resultatet från mätningarna användas för att
matematiskt uppskatta det dagliga intaget. I studien utvärderades även
förekomsten av dessa mjukgörare genom att bestämma halten av deras
metaboliter i prover av fingernaglar.
Då referensstandarder för alternativa mjukgörare, undantaget DINCH, inte är
kommersiellt tillgängliga utfördes in vitro experiment för att kartlägga
huvudmetaboliterna till fyra alternativa mjukgörare (så som DEHTP, di(2-
propylheptyl) ftalat (DPHP), di(2-etylhexyl) adipat (DEHA) och acetyl tributyl
citrat (ATBC)). Metaboliterna identifierades med hjälp av en högupplösande
masspektrometer kopplat till en vätskekromatograf (Q Exactive Orbitrap LC-
MS). De fastställda metaboliterna från in vitro försöken användes som
biomarkörer för att uppskatta mänsklig exponering för dessa kemikalier genom
att bestämma deras förekomst i prover från naglar och urin. Då många av de
metaboliter från alternativa mjukgörare som påträffades i prover in vitro
återfanns även in vivo i urin- och nagelprover, är slutsatsen att resultat från in
vitro försök gällande mjukgörare i plast uppvisar god överensstämmelse med
verklig metabolism. Resultatet av studien visar även att fingernaglar som matris
för human biomonitorering kan utgöra en andvändbar icke-invasiv provtagning
för bestämning av humanexponering av organiska miljöföroreningar, men att
behovet av ytterligare validering behövs för att dra långtgående slutsatser.
I studien bestämdes även förekomsten av PEs och DINCH i dubbelprover
från kost, inandningsluft, damm och handavtork. Extern exponering, uppskattad
utifrån upptag via kost, luft, och hud, visade på resultat som låg högre eller lika
med det matematiskt uppskattade dagliga intaget. Genom att jämföra de olika
exponeringsvägarna var det möjligt att dra slutsatser kring den relativa
betydelsen av olika exponeringsvägar för den Norska populationen. Slutsatsen är
att intag av dessa ämnen via kosten utgör merparten av det som återfinns i
kroppen. upptag via inandning har endast betydelse för låg molekylära PE,
medan intag av dammpartiklar förklarar mänskligt upptag av hög molekylära
PE, inklusive DINCH. Upptag via huden och handavtork låg mycket lägre än
det samantagna upptaget som uppskattats via luft, damm och
personvårdsprodukter, men fortfarande råder många osäkerheter kring denna
exponeringsväg.
Baserat på beräkningar av intaget överskred inte de norska deltagarna i
studien tolerabelt dagligt intag, även om summativa kombinations effekter av
andra förekommande hormonstörande ämnen togs i beaktan. Dock visar studien
att det är av yttersta vikt att kontinuerligt monitorera dessa alternativa
mjukgörare då användningen ständigt ökar.
Resultat av multivariata analyser från studien visar att faktorer, såsom ålder,
rökning, och användning av hygienprodukter har betydelse. Men även att rutiner
såsom frekvensen av att tvätta händerna, äta snabbmat med händerna eller att äta
mat ur en burk i plast utgör möjliga vägar för mänsklig exponering av dessa
kemikalier.
Nyckelord: Ftalater, Alternativa mjukgörare, DINCH, In vivo screening, Urin,
Naglar, Luft, Damm, Handavtork, Mat, Prediktorer för human exponering
List of papers
The thesis is based upon the following papers, referred to in the text by their
Roman numerals (I-IV).
Paper I Human exposure, hazard and risk of alternative plasticizers to phthalate
esters.
Bui TT, Giovanoulis G, Cousins AP, Magnér J, Cousins IT, de Wit CA.
Sciences of Total Environment. 2016 Jan 15;541:451-67. Review.
doi: 10.1016/j.scitotenv.2015.09.036.
Reproduced with permission from Elsevier
Paper II Evaluation of exposure to phthalate esters and DINCH in urine and nails
from a Norwegian study population.
Giovanoulis G, Alves A, Papadopoulou E, Cousins AP, Schütze A, Koch
HM, Haug LS, Covaci A, Magnér J, Voorspoels S.
Environmental Research. 2016 Nov;151:80-90.
doi: 10.1016/j.envres.2016.07.025
Reproduced with permission from Elsevier
Paper III Case study on screening emerging pollutants in urine and nails.
Alves A, Giovanoulis G, Nilsson UL, Erratico C, Lucattini L, Haug LS,
Jacobs G, de Wit CA, Leonards PE, Covaci A, Magnér J, Voorspoels S.
Environmental Sciences & Technology. 2017 Apr 4;51(7):4046-4053.
doi: 10.1021/acs.est.6b05661
Reproduced with permission from American Chemical Society
Paper IV Multi-pathway human exposure assessment of phthalate esters and
alternative plasticizers.
Giovanoulis G, Bui T, Xu F, Covaci A, Haug LS, Cousins AP, Cousins IT,
Magnér J, de Wit CA
Submitted to Environment International
Statement of contribution
I (Georgios Giovanoulis) shared the first co-authorship in all four papers
presented in this thesis.
Paper I I was involved in planning of the paper together with other co-authors. I
contributed by collecting literature information and together with T. Bui,
took the lead role in writing the paper. A.P. Cousins, J. Magnér, I.T.
Cousins and C.A. de Wit read and commented on the manuscript.
Paper II I was involved in developing the sampling protocols and participated in the
sampling campaign. Together with A. Alves, I performed all the
experimental work, including laboratory work, instrumental and data
analysis. Statistical analysis and risk assessment were carried out with the
help of E. Papadopoulou and A.P. Cousins, respectively. I took the lead
role in writing the paper. All co-authors read and commented on the
manuscript.
Paper III I was involved in developing the idea and the experimental design. I was
equally responsible with A. Alves for all the experimental work including
laboratory work, instrumental and data analysis. The in vitro metabolite
standards were prepared by C. Erratico, A. Alves and L. Lucattini. The
interpretation of the fragments using mass spectrometry was done with the
help of U. Nilsson. I participated in writing the paper. All co-authors read
and commented on the manuscript.
Paper IV I was involved in developing the sampling protocols and participated in the
sampling campaign. I was responsible for all the experimental work,
including laboratory work, instrumental and data analysis. Together with T.
Bui, I carried out the statistical analysis, risk assessment and writing of the
paper. All co-authors read and commented on the manuscript.
Other related work
The following papers are related to my PhD studies but they are not included
in this thesis.
Paper V Comprehensive Study of Human External Exposure to Organophosphate
Flame Retardants via Air, Dust, and Hand Wipes: The Importance of
Sampling and Assessment Strategy.
Xu F, Giovanoulis G, van Waes S, Padilla-Sanchez JA, Papadopoulou E,
Magnér J, Haug LS, Neels H, Covaci A.
Environmental Sciences & Technology. 2016 Jul 19;50(14):7752-60.
doi: 10.1021/acs.est.6b00246
Paper VI Phthalates, non-phthalate plasticizers and bisphenols in Swedish
preschool dust in relation to children's exposure.
Larsson K, Lindh CH, Jönsson BA, Giovanoulis G, Bibi M, Bottai M,
Bergström A, Berglund M.
Environment International. 2017 Mar 5. pii: S0160-4120(16)30792-9.
Open access.
doi: 10.1016/j.envint.2017.02.006
Paper VII Mass transfer of an organophosphate flame retardant between product
source and dust in direct contact.
Liagkouridis I, Lazarov B, Giovanoulis G, Cousins IT.
Submitted to Chemosphere.
Paper VIII In vitro inhalation bioaccessibility of plasticizers present in indoor dust
using artificial lung fluids.
Kademoglou K, Giovanoulis G, Cousins AP, de Wit CA, Haug LS,
Williams AC, Magnér J, Collins CD.
Manuscript
Table of contents
1 Introduction ........................................................................................................... 1
1.1 Phthalates & alternatives ................................................................................. 1
1.2 Plasticizers in the human body ........................................................................ 2
1.3 Human exposure to plasticizers ...................................................................... 3
1.4 Exposure pathways ......................................................................................... 4
1.5 Internal exposure ........................................................................................... 11
1.6 External exposure .......................................................................................... 13
2 Objectives ............................................................................................................. 16
3 Materials & methods ........................................................................................... 18
3.1 Advanced Tools for Exposure Assessment & Biomonitoring (A-TEAM) ... 18
3.2 Sampling campaign ....................................................................................... 18
3.3 Sample treatment & instrumental analysis .................................................... 19
3.3.1 Precautions against contamination ....................................................... 19
3.3.2 Urine & nail analysis ............................................................................ 20
3.3.3 In vivo screening .................................................................................. 21
3.3.4 Air, dust, hand wipe & food analysis ................................................... 21
3.4 Quality control .............................................................................................. 22
3.5 Human exposure estimation & risk assessment ............................................ 23
3.6 Potential predictors of exposure .................................................................... 25
4 Results & discussion ............................................................................................ 27
4.1 Alternative plasticizers to phthalate esters .................................................... 27
4.2 Urine & nails ................................................................................................. 30
4.3 Potential predictors of exposure .................................................................... 31
4.4 Air, dust, hand wipes & food ........................................................................ 32
4.5 Human exposure estimation & risk assessment ............................................ 33
4.6 Relative importance of different exposure pathways .................................... 36
4.7 In vivo screening ........................................................................................... 36
5 Concluding remarks ............................................................................................ 39
6 Future perspectives .............................................................................................. 41
7 Acknowledgments ................................................................................................ 44
8 References ............................................................................................................ 45
Abbreviations
A-TEAM advanced tools for exposure assessment and biomonitoring
ATBC acetyl tributyl citrate
BBzP benzyl butyl phthalate
DMP dimethyl phthalate
DEP diethyl phthalate
DnBP di-n-butyl phthalate
DiBP diisobutyl phthalate
DEHP di(2-ethylhexyl) phthalate
DiNP diisononyl phthalate
DiDP diisodecyl phthalate
DPHP di(2-propylheptyl) phthalate
DINCH cyclohexane-1,2-dicarboxylic acid diisononyl ester
DEHTP di(2-ethylhexyl) terephthalate
DEHA di(2-ethyhexyl) adipate
DF % detection frequency
DI daily intake
ESBO epoxidized soybean oil
ESI electrospray ionization
FR flame retardant
FUE urinary molar excretion factors
GC gas chromatography
GTA glycerin triacetate
HBM human biomonitoring
HI hazard index
HPLC high pressure liquid chromatography
HPV high production volume
HQ hazard quotient
KOA octanol-air partition coefficient
KOW octanol-water partition coefficient
MAE microwave assisted extraction
MRM multiple reaction monitoring
MS mass spectrometer
MEHTP mono(2-ethylhexyl) terephthalate
OH-MEHTP 1-mono(2-ethylhydroxyhexyl) benzene-1,4-dicarboxylate
PE phthalate ester
PVC polyvinyl chloride
PCPs personal care products
POP persistent organic pollutant
PBT persistent bioaccumulative and toxic
PK pharmacokinetic
QC quality control
REACH registration evaluation authorization and restriction of chemicals
RfD reference dose
SMURF Stockholm multimedia urban fate
SPE solid phase extraction
SRM standard reference material
SI supporting information
TCP tricresyl phosphate
TEHPA tris(2-ethylhexyl) phosphate
TDI tolerable daily intake
TXIB trimethyl pentanyl diisobutyrate
V6 2,2-bis(chloromethyl)-propane-1,3-diyltetrakis(2-chloroethyl)
bisphosphate
vPvB very persistent, very bioaccumulative
1
1 Introduction
1.1 Phthalates & alternatives
Modern life is inseparable from plastic materials. The enhanced elasticity
and durability of these polymeric products is due to phthalate esters (PEs), a
group of synthetic chemicals used as plasticizers and additives. Several million
tons of PEs are produced worldwide each year for the production of soft
polyvinyl chlorine (PVC) (Jaakkola and Knight 2008). Typical products
containing plasticizers are PVC resins, floorings, wall coverings, cables, textiles,
children’s toys, personal care products (PCPs), food package materials and
medical devices (e.g. infusion tubings, blood bags, etc.) (Wormuth et al. 2006).
The most important PEs (Table 1) can be classified according to their chemical
structure as low molecular weight: dimethyl phthalate (DMP), diethyl phthalate
(DEP), di-n-butyl phthalate (DnBP), diisobutyl phthalate (DiBP) and
butylbenzyl phthalate (BBzP), and high molecular weight: di(2-ethylhexyl)
phthalate (DEHP), diisononyl phthalate (DiNP), diisodecyl phthalate (DiDP)
and di(2-propylheptyl) phthalate (DPHP).
The interaction of plasticizers with the polymer resin macromolecules is not
permanent since they are not covalently bound. Consequently, plasticizers can
be released and contaminate the environment by migration, evaporation,
leaching and abrasion from the products (Wittassek et al. 2011). The extensive
use of PEs has caused these contaminants to be ubiquitous in environments
relevant for human exposure, especially indoors where the levels are
comparatively higher, and as a result also in human tissues (Koch and Calafat
2009; Wittassek and Angerer 2008; Wittassek et al. 2011).
Endocrine disruption, developmental and reproductive toxicity are common
observed effects in humans and animals for certain PEs, such as DiBP, DnBP,
BBzP and DEHP (Foster et al. 2001; Ventrice et al. 2013; Wittassek and
Angerer 2008). For these four chemicals a special authorization is required for
any kind of application (EC 2006), while there is a proposal to further restrict
their aggregate use in articles (< 0.1% w/w of the plasticized material) on the
2
EU market (ECHA 2016). Moreover, children’s exposure to PEs has been linked
to increased risk of allergies and asthma (Braun et al. 2013; Sathyanarayana
2008). Infants, children and pregnant women are sensitive population subgroups
since they are more susceptible to PE toxicity (Heudorf et al. 2007). Therefore,
DiNP and DiDP are also prohibited in toys which can be mouthed by children
(EC 2005). In certain consumer product applications there is a lack of strict
regulation. For instance DEP is still widely used in PCPs, even though it is
suspected to increase the breast cancer risk (Lopez-Carrillo et al. 2010).
However, actions to alert populations have arisen, e.g. the Campaign for Safe
Cosmetics (www.safecosmetics.org), and resulted in reduction of DEP exposure
over the last decade (Giovanoulis et al. 2016; Zota et al. 2014).
As a consequence, many different alternative plasticizers to PEs (Table 1)
have been introduced onto the market for applications with close human contact.
Cyclohexane-1,2-dicarboxylic acid diisononyl ester (DINCH) is a less toxic,
non-aromatic replacement, mainly for DEHP and DiNP, in children’s toys, food
packaging and medical devices, with less potential for leaching (EFSA 2006;
Holmgren et al. 2012; Zhong et al. 2013). Di(2-ethylhexyl) terephthalate
(DEHTP) is a substitute for DEHP in plastic toys and childcare articles, films,
pavement, stripping compounds, walk-off mats, vinyl products and beverage
closures, and it has two adjacent ring substitutions occupying para-positions
instead of ortho-positions of DEHP (Bui et al. 2016; Eastman 2011; Lioy et al.
2015). Acetyltributylcitrate (ATBC) is used in food packaging, medical products
and PCPs (Johnson 2002; Stuer-Lauridsen et al. 2001), while di(2-ethyhexyl)
adipate (DEHA) is used mainly in applications intended for food wrapping.
1.2 Plasticizers in the human body
Elimination of PEs inside the human body is fast (24-48 h). In the phase I
metabolic step, parent diesters entering the human body are rapidly metabolized
to their respective primary/hydrolytic monoesters. In the case of high molecular
weight PEs and alternative plasticizers (e.g. DINCH, DEHTP and DEHA), the
3
hydrolytic monoesters are further metabolized to various secondary/oxidative
metabolites (Koch et al. 2005; Koch et al. 2013; Lessmann et al. 2016).
Hydrolytic and oxidative (phase I) monoester metabolites can be excreted by
humans mainly through urine (and partly with feces and sweat) unchanged, or
they can undergo phase II biotransformation to produce glucuronide conjugates
with higher water solubility, facilitating excretion (Calafat et al. 2006). The ratio
between free monoester and glucuronide conjugate excretion varies among
different PEs (Hauser and Calafat 2005).
Despite the fast elimination and excretion of PEs by humans, they are
pseudo-persistent due to their continuous production and release into the
environment, leading to continuous environmental and human exposure (Bui et
al. 2016; Mackay et al. 2014).
1.3 Human exposure to plasticizers
External exposure expresses the concentration of plasticizers released from a
source of pollution which comes into contact with humans for a certain period of
time (acute, short-term, chronic and life-long exposure). Internal exposure to
plasticizers is the amount of parent diesters that enter the human body and are
metabolized, which indicates the total human body burden of the exposure. The
internal exposure can be measured using urinary metabolite concentrations and
their urinary molar excretion factors (FUE) presented in Table 1. There are
certain factors that might influence external and internal exposure, such as
physiological factors (age, gender, body weight, skin surface area, nutritional
status, disease and genetics) and exposure factors related to human behavior and
activities (time-activity patterns, life-style factors, socioeconomic status and
physical activity). The best way to calculate or estimate the magnitude,
frequency, and duration of plasticizer exposure, along with the number and
characteristics of the population exposed is to perform a human exposure
assessment. Ideally, it describes the sources, pathways, routes, and the
4
uncertainties in the assessment (IPCS 2004). Different approaches to human
exposure assessment are presented in Figure 1.
Figure 1. Different approaches to human exposure assessment.
1.4 Exposure pathways
Human exposure to plasticizers mainly occurs through ingestion, inhalation,
and dermal uptake (Clark et al. 2011; Wormuth et al. 2006). Also, sometimes
intravenous injection can be a route of human exposure through a variety of
PVC medical devices (Calafat et al. 2004). Air, dust, PCPs and diet are the most
relevant sources of plasticizer exposure.
Excluding dietary intake, inhalation of contaminated air has been confirmed
as the predominant pathway of external exposure to more volatile PEs (Wang et
al. 2013). The concentrations in indoor air are generally higher than outdoor
concentrations (Rudel and Perovich 2009). Low molecular weight PEs are
ubiquitous in indoor air (Bergh et al. 2011a; Bergh et al. 2011b) while the high
molecular weight PEs are more prevalent in house dust (Heudorf et al. 2007).
PE exposure from unintended dust ingestion, dust adhered to the skin and
inhalable dust (<100 μm) may have negative impacts on human health. Dust is
an important source of children’s exposure due to their close-to-floor activities
and frequent hand-to-mouth contact (Larsson et al. 2017). Several studies have
5
measured PEs in dust from micro-environments relevant to children, such as
daycare centers and preschools (Gaspar et al. 2014; Langer et al. 2010;
Myridakis et al. 2016). Also, alternative plasticizers to PEs, especially DINCH,
have been found in considerable amounts in these micro-environments (Fromme
et al. 2016; Larsson et al. 2017).
A variety of cosmetics and PCPs has been identified as sources of
transdermal exposure to certain PEs, such as DEP (Guo and Kannan 2013;
Koniecki et al. 2011; Koo and Lee 2004; Wormuth et al. 2006). Women have a
significantly higher risk of exposure to DEP than men due to more frequent use
of PCPs (Romero-Franco et al. 2011; Wittassek et al. 2011). Furthermore, the
direct transdermal uptake from air is not routinely considered, and has only
recently been included in human exposure assessments (Beko et al. 2013;
Gaspar et al. 2014).
Diet is the predominant source of PE exposure (Wormuth et al. 2006) due to
partitioning of these chemicals into food, water and beverages from packaging
materials during processing and storage. Also, cooking equipment (e.g. plastic
handles and cutting boards) and PVC gloves for food preparation might
introduce contamination. Finally, breast milk can be a medium of concern for
early life stage exposure to PEs. Considering that breast milk is often the only
nutritional source for newborn infants, efforts to reduce PE exposure among
lactating women are required (Kim et al. 2015).
Although the exposure pathways to “classical” PEs are well studied there is
lack of data for the new alternative plasticizers. Only a few studies have
analyzed these chemicals in different external exposure media (Cao et al. 2013;
Fromme et al. 2013; Fromme et al. 2016; Larsson et al. 2017). The importance
for inclusion of new alternative plasticizers in future exposure studies is also
supported by model calculations for the release of plasticizers from PVC
flooring used in Sweden. These showed reduced DEHP emissions due to
substitution with DINP, and that in the near future DINCH would yield similar
emissions to DINP (Holmgren et al. 2012).
6
Table 1. Names, abbreviations, CAS registration numbers, chemical structures and physicochemical properties of phthalate
esters and their main alternative plasticizers. Primary (hydrolytic monoester) and secondary (oxidized monoester)
metabolites are illustrated together with the human urinary molar excretion fractions (FUE) for each metabolite as a
percentage of the ingested dose of the parent compound at 24 h post exposure.
Name Chemical structure MW [g/mol] Water solubility
[mg/l] 25°C
Primary/hydrolytic
metabolites FUE
[%]
Referenc
es (Abbreviation) CAS number
Vapor
pressure [Pa]
25°C
log KOW 25°C Secondary/oxidative
metabolites
Ph
tha
late
est
ers
Dimethyl
phthalate
(DMP) 131-11-3
194.18
2.39 a
2.19 × 10-2 a
1.53 a
mono-methyl phthalate
(MMP) 69 e
(Koch and
Angerer
2012)
Diethyl phthalate
(DEP) 84-66-2
222.24
0.22 a
3.63 × 10-3 a
2.39 a
mono-ethyl phthalate
(MEP) 69
(Koch
and
Angerer 2012)
Diisobutyl
phthalate
(DiBP) 84-69-5
278.34
0.005 a
4.37 × 10-5 a
4.61 a
mono-isobutyl phthalate (MiBP)
70 (Koch et al. 2012)
Di-n-butyl
phthalate
(DnBP) 84-74-2
278.34
0.005 a
4.37 × 10-5 a
4.61 a
mono-n-butyl phthalate
(MnBP) 84
(Koch et
al. 2012)
7
Benzyl butyl phthalate
(BBzP) 85-68-7
312.37
0.001 a
8.32 × 10-6 a
4.91 a
mono-benzyl phthalate
(MBzP) 73
(Anderson et al.
2001)
Di(2-
ethylhexyl)
phthalate
(DEHP) 117-81-7
390.56
7.59 × 10−4 a
7.41 × 10-8 a
7.48 a
mono-(2-ethylhexyl)
phthalate (MEHP)
5.9
(Koch et
al. 2005)
mono(2-ethyl-5-
hydroxyhexyl) phthalate
(5-OH-MEHP) mono(2-ethyl-5-oxohexyl)
phthalate
(5-oxo-MEHP) mono(5-carboxy-2-
ethylpentyl) phthalate
(5-cx-MEPP) mono(2-
carboxymethylhexyl)
phthalate (2-cx-MEPP)
23.3
15
18.5
4.2
Diisononyl
phthalate
(DiNP)
28553-12-0
418.61
5.17 × 10−6 c
1.74 × 10−5 b
9.52 b
mono-isononyl phthalate (MiNP)
2.12
(Koch
and Angerer
2007)
mono(4-methyl-7-hydroxyoctyl) phthalate
(7OH-MMeOP)
mono(4-methyl-7-oxo-octyl) phthalate
(7oxo-MMeOP)
mono(4-methyl-7-carboxyheptyl) phthalate
(7cx-MMEHP)
18.4
9.97
9.07
8
Diisodecyl
phthalate
(DiDP) 26761-40-0
446.74
1.84 × 10−6 c
2.24 × 10−6 b
9.46 b
mono-iso-decyl phthalate (MiDP)
2.12 f
(Silva et
al.
2007a)
mono-hydroxyl-iso-decyl
phthalate (OH-MIDP)
mono-oxo-iso-decyl
phthalate (oxo-MIDP)
mono-carboxy-iso-nonyl
phthalate (cx-MIDP)
18.4 f
9.97 f
9.07 f
Di(2-propylheptyl)
phthalate
(DPHP) 53306-54-0
446.66
4.91 × 10−6 b
2.2 × 10−6 b
10.36 b
mono-propylheptyl phthalate
(MPHP) <1
(Gries et al. 2012;
Leng et
al. 2014)
mono(2-propyl-6-hydroxy-
heptyl) phthalate (OH-MPHP)
mono(2-propyl-6-oxo-heptyl)
phthalate (oxo-MPHP)
mono(2-propyl- 6-carboxy-
hexyl) phthalate (cx-MPHxP)
9.91
12.61
0.42
Alt
ern
ati
ve
pla
stic
izers
Di(2-
ethylhexyl)
adipate
(DEHA)
130-23-1
370.57
4.27 × 10−4 d
5.45 × 10−3 d
8.94 d
mono(2-ethylhexyl) adipate
(MEHA) -
(Silva et al.
2013b)
mono(2-ethylhydroxyhexyl)
adipate
(MEHHA) mono(2-ethyloxohexyl)
adipate
(MEOHA)
-
-
9
Acetyltributylcitrate
(ATBC)
77-90-7
402.48
6.07 × 10−4 d
0.65 d
4.29 d
acetyl citrate
monobutyl citrate acetyl monobutyl citrate
dibutylcitrate acetyldibutyl
citrate
-
-
- -
(Johnson
2002)
Di(2-
ethylhexyl) terephthalate
(DEHTP)
6422-86-2
390.56
2.86 × 10−3 d
2.39 × 10−4 d
8.39 d
mono(2-ethylhexyl)
terephthalate (MEHTP)
-
(Lessman
n et al.
2016)
1-mono(2-ethyl-5-hydroxy-
hexyl)benzene-1,4-
dicarboxylate (5OH-MEHTP)
1-mono(2-ethyl-5-oxo-
hexyl)benzene-1,4-dicarboxylate
(5oxo-MEHTP)
1-mono(2-ethyl-5-carboxyl-pentyl)benzene-1,4-
dicarboxylate (5cx-MEPTP)
1-mono(2-carboxyl-methyl-
hexyl)benzene-1,4-
dicarboxylate
(2cx-MMHTP)
1.72
0.95
12.24
0.27
Cyclohexane-
1,2-dicarboxylic
acid
diisononyl ester
(DINCH)
166412-78-8
424.65
1.28 × 10−4 d
8.8 × 10−6 d
10 d
mono-iso-nonyl-
cyclohexane-1,2-dicarboxylate
(MINCH)
0.65
(Koch et al. 2013)
cyclohexane-1,2-dicarboxylic mono hydroxyisononyl ester
(OH-MINCH)
9.55
10
cyclohexane-1,2-dicarboxylic
mono oxoisononyl ester (oxo-MINCH)
cyclohexane-1,2-diarboxylic
mono carboxyisononyl ester (cx-MINCH)
1.85
1.67
a Schwarzenbach (2005), b estimated using EPISuite Version 4.1, c Yaws et al. (1994), d Bui et al. (2016), e Analogously to DEP, f Analogously to DiNP
11
1.5 Internal exposure
Human biomonitoring (HBM) is a commonly used technique to measure the
internal exposure by assessing whether and to what extent chemicals enter the
human body. It is ideal for risk assessment and risk management since it
considers all relevant routes and sources of human exposure. In the case of PEs,
blood and urine are traditionally the most widely analyzed matrices for
determination of internal exposure (Angerer et al. 2007). Other non-invasive
matrices, such as hair, nails and saliva have been introduced in the HBM field,
due to advantages with respect to cost reduction of sampling procedures, less
storage space, sample stability and possible simplification of the ethical
approval and recruitment (Alves et al. 2014; Smolders et al. 2009).
An important consideration in HBM is the proper matrix selection for the
biomonitoring purpose. For instance, urine, serum, and saliva from nursing
mothers cannot be used to estimate exposure to PEs for breastfeeding infants
(Hogberg et al. 2008). In particular, urinary PE concentrations reflect maternal
exposure, but they do not represent the metabolite concentrations in other
biological fluids, especially breast milk (Hines et al. 2009).
Measurements of PE metabolites in blood, saliva and breast milk might be
hampered by external contamination of samples with PEs (during sample
collection, storage and/or analysis) that can be hydrolyzed to their respective
monoesters by esterases. Therefore, if the enzymatic activity in the matrix is not
eliminated, phthalate monoester measurements will be artificially high. It is
possible to prevent immediate breakdown of the parent diester compound to the
monoester metabolites by adding an esterase inhibitor (e.g. phosphoric acid)
during the sample collection, and directly storing the samples at -20 ᴼC (Hines et
al. 2009; Hogberg et al. 2008; LaKind et al. 2009).
To assess human exposure to non-persistent chemicals like PEs, urine is
generally the matrix of choice because metabolite levels in urine are more
informative and much higher than in blood (Alves et al. 2014; Frederiksen et al.
2010). In urine, it is possible to back-calculate the initial exposure to parent
plasticizers by using metabolite concentrations and their excretion factors (Table
12
1). Primary/hydrolytic monoesters are useful biomarkers for human exposure
assessment, especially for low molecular weight PEs (no secondary metabolic
oxidation occurs). As mentioned above, hydrolytic monoesters are susceptible to
external contamination to parent diesters (Wittassek and Angerer 2008) or they
can be present in trace levels in milk and dairy products due to animal
metabolism. However, overestimation of phthalate monoesters in biological
fluids (i.e. human milk, blood and urine) because of their initial existence in
food is considered to be negligible (Cao 2010). Moreover, in the case of high
molecular weight PEs hydrolytic monoesters are only minor metabolites in urine
and have shorter half-lives (Koch et al. 2005; Preuss et al. 2005; Wittassek and
Angerer 2008). Therefore, appropriate biomarkers of exposure are essential for
urinary metabolite analysis in HBM studies. The secondary oxidized (OH-, oxo-
and carboxy (cx-)) metabolites of PEs are sensitive and specific biomarkers of
exposure, and they are preferred whenever possible. These metabolites are
unique, independent of external contamination, and can be formed only by
oxidation of monoester metabolites (Koch et al. 2005). Attempts to quantify the
total exposure of high molecular weight PEs without measuring specific
metabolites will result in underestimation.
Metabolism of commonly used high molecular weight PEs, such as DEHP,
DiNP and DiDP is well understood (Koch and Angerer 2007; Koch et al. 2004;
Silva et al. 2007a) and their specific biomarkers of exposure have been widely
used in several HBM studies to determine the internal exposure (Guo et al.
2011; Koch et al. 2011; Larsson et al. 2014; Silva et al. 2007b; Wittassek et al.
2007b). Recently, accurate urinary excretion factors of specific biomarkers were
derived for the new alternative plasticizers DPHP, DINCH and DEHTP (Koch et
al. 2013; Leng et al. 2014; Lessmann et al. 2016) and has allowed their inclusion
in biomonitoring of different study populations (Correia-Sá et al. 2017; Fromme
et al. 2016; Gomez Ramos et al. 2016; Hauser et al. 2016; Lessmann et al. 2017;
Schutze et al. 2014; Silva et al. 2013a).
In HBM assessments, the use of unconventional matrices is becoming an
increasingly important area of research, for example the use of hair and nails.
13
Apart from being non-invasive, easier to collect and store, and inexpensive, they
can provide a longer detection window (usually months to years), enabling
retrospective estimation of chronic and past exposure, which is not always
possible for blood, serum, plasma, or urine samples (Chang et al. 2013).
However, analytical challenges exist in application of non-invasive matrices.
These include low concentrations of targeted chemicals which makes it difficult
to perform real human assessment of body burdens. There have been only a few
biomonitoring studies that included the most accessible non-invasive matrices,
such as hair, saliva and nails, as biomarkers of human exposure to PEs (Alves et
al. 2016a; Chang et al. 2013; Hines et al. 2009; Silva et al. 2005).
HBM data can be used to calibrate pharmacokinetic models that predict
concentrations in different body compartments. An empirical pharmacokinetic
(PK) model has been developed to predict DiBP, DnBP and DEHP metabolite
levels in urine and serum after oral doses of the parent compounds (Lorber et al.
2010; Lorber and Koch 2013). Urinary metabolite data from five individuals in a
fasted state over 48 h were used to validate this PK model, showing a good fit
for most of the metabolites. Also, a multi compartment PK model can be used to
characterize the exposure to DINCH (Schutze et al. 2015). This model was
calibrated using urine metabolite levels from three different male volunteers,
each orally dosed with 50 mg DINCH. The predicted values showed a good
agreement with the observed urinary DINCH metabolite concentrations. Finally,
Bui et al. (2017) recently provided insights into the partitioning of PE
metabolites between blood and nails using PK modelling (Lorber et al. 2010)
and biomonitoring data from the Norwegian study population included in this
thesis (paper II).
1.6 External exposure
External exposure estimated through environmental and personal monitoring
(i.e. air, dust, hand wipes and diet) can be used to elucidate the sources of
exposure and the relative importance of the various exposure pathways.
14
However, to confirm whether an exposure pattern for chemicals is well
understood or not, human intake based on external exposure estimates should be
compiled and compared with estimates based on body burden (i.e. internal
concentrations). An agreement between these two estimates indicates a good
understanding of the relevant sources and pathways of exposure. An attempt to
verify this for plasticizers should ideally measure the parent compounds in
external exposure media, as well as biomarkers in urine.
Wormuth et al. (2006) investigated in detail European consumers’ daily
exposure to eight frequently used PEs based on external exposure media, and
tried to link the knowledge on emission sources of PEs and the concentrations of
PE metabolites found in urine. Other studies have also estimated daily intakes
(DIs) of PEs using both external and biomarker based methods, and usually the
two approaches agreed with each other within an order of magnitude (Clark et
al. 2011; Itoh et al. 2007). Lack of information concerning human absorption of
different PEs after oral exposure, regional differences and temporal changes in
the use of the PEs might introduce discrepancies between the two approaches
(Clark et al. 2011). Finally, Beko et al. (2013) performed an intake estimation of
several PEs (i.e. DEP, DnBP, DiBP, BBzP and DEHP) for children and found
that their estimations based on urinary metabolite levels agree with those in
other studies (Frederiksen et al. 2011; Koch et al. 2007; Wittassek et al. 2007a).
However, their estimates from external exposure (based on dust ingestion,
inhalation and dermal absorption) could explain 100 % of the total intake only
for DEP. For other compounds, knowledge gaps obviously exist as the
calculations from external exposure could only explain about 50 %, 17 %, 8 %
and 3 % for DiBP, DnBP, DEHP and BBzP, respectively.
As mentioned above (section 1.3), oral exposure via dietary intake seems to
constitute the major source of plasticizer exposure (Wormuth et al. 2006). Dust
ingestion, to some extent, might also contribute, especially for children (Larsson
et al. 2017). Inhalation of air, transdermal exposure through the use of PCPs and
dust adhered to the skin may also make some contribution (Beko et al. 2013;
Koniecki et al. 2011; Romero-Franco et al. 2011). Other routes of exposure,
15
such as direct dermal contact with plastic materials, air-to-skin transdermal
uptake and hand-to-mouth contact are not well investigated and data gaps
remain.
16
2 Objectives
The main objective of this thesis was to perform a comprehensive evaluation
of exposure to PEs and new alternative plasticizers for the included Norwegian
population. The following questions on human exposure to PEs were addressed:
1. What is the state of knowledge regarding alternative plasticizers to PEs
and what are the major data gaps? (paper I)
2. What is the daily intake of plasticizers based on the internal exposure and
is it possible to link the exposure with sociodemographic and lifestyle
characteristics? (paper II)
3. Are nails a possible alternative non-invasive matrix for estimating human
exposure to plasticizers? (papers II & III)
4. Can in vitro assays reliably mimic in vivo metabolic processes? (paper
III)
5. How comparable are external and internal exposure estimates for PEs and
DINCH? (paper IV)
6. Are all pathways of exposure well understood and what is their relative
importance in contributing to body burdens? (paper IV)
Paper I aimed to evaluate current substance classes of alternative
plasticizers to PEs including discussion of physicochemical properties,
production volumes, use, emissions, indoor fate, human exposure and health
concerns, and by addressing their human risk potential and their persistent /
bioaccumulative / toxic (PBT) properties. The main objective was to identify
key data gaps for more comprehensive risk assessment.
Paper II aimed to evaluate the internal exposure and perform a cumulative
risk assessment of the Norwegian study population to PEs and DINCH. This
study also aimed to assess the correlations between urinary and nail metabolite
17
concentrations, their relationships with different sociodemographic and lifestyle
characteristics, and the applicability of nails as a non-invasive matrix in human
exposure assessment.
Reference standards of human metabolites for many emerging pollutants are
still not commercially available and their positive identification in vivo becomes
rather difficult. It is crucial to discover specific metabolites which can then be
used as appropriate biomarkers for human exposure to these chemicals. The
main objective in paper III was to identify specific metabolites which can serve
as new biomarkers for human exposure to emerging pollutants (i.e. DPHP,
DEHTP, ATBC, DEHA and V6). Another aim was to screen urine and finger
nails in vivo for the major phase I metabolites, which had been previously
identified in in vitro assays, to test the validity of using urine and nails as a non-
invasive method to assess exposure to new pollutants.
Paper IV aimed to determine external exposure to PEs and DINCH from
house dust, personal and indoor air, diet and personal (hand wipes) samples
from the Norwegian study population. Other objectives of this study were to
evaluate all external routes of exposure and elucidate their relative importance,
compare DI estimates based on external and internal exposure and perform an
exposure and risk assessment for the Norwegian cohort.
18
3 Materials & Methods
3.1 Advanced Tools for Exposure Assessment &
Biomonitoring (A-TEAM)
This thesis was part of the A-TEAM project, which aimed to increase
knowledge for a variety of aspects related to internal and external human
exposure to selected consumer chemicals (PEs; organophosphate esters;
perfluoroalkyl substances; brominated flame retardants). The enhanced
understanding of the underpinning science could benefit scientists, public health
and associated regulators by delivering more effective approaches to monitoring
human exposure to chemicals within Europe. This thesis is focused on PEs and
alternative plasticizers.
3.2 Sampling campaign
The study population consisted of 61 adults (age: 20–66; gender: 16 males
and 45 females) living in the Oslo area (Norway). Sampling was conducted
during the winter of 2013–2014, and indoor environment, dietary and biological
samples were collected from each individual participant and their household,
during a 24 h period. Information about the home environment, personal and
lifestyle characteristics was collected via questionnaires (Papadopoulou et al.
2016).
Three urine spots (afternoon – day 1, morning – day 2 and afternoon – day 2)
and fingernails were collected from each participant and used to study internal
exposure. Participants were asked not to cut their fingernails for 2–3 weeks prior
to the sample collection, and they were advised to remove any nail polish, dirt,
debris and artificial nails before clipping their fingernails.
Air, dust, hand wipes, and food from a duplicate 24 h diet were collected to
study different routes of external exposure. Indoor air sampling was performed
by placing a stationary SKC Leland Legacy pump (SKC Inc., Pennsylvania,
U.S.) connected to four, in parallel, ENV+ solid phase extraction (SPE)
19
cartridges (200 mg, 6 mL, Biotage, Uppsala, Sweden) in the living room for 24
h. At the same time, a personal ambient air sample was collected by using a
portable SKC 224-PCMTX4 pump (SKC Inc., Pennsylvania, U.S.) connected to
one ENV+ SPE cartridge (1 g, 25 mL, Biotage Uppsala, Sweden) close to the
participant’s face during the entire 24 h sampling event, including sleeping
hours. Dust from the floor and elevated surfaces in the living room were
collected separately from each house using a vacuum cleaner equipped with a
forensic nozzle and a one-way filter housing a dust-sampling filter. Participants
were also asked not to discard their vacuum cleaner bags and to provide it to the
researchers at the end of the sampling event. The food samples consisted of
homogenized duplicate portions of all foods consumed over 24 h. Hand wipes
were collected from both hands of the participants, who were recommended not
to wash their hands 60 min prior to the collection. Each hand (palm and back-of-
hand) was thoroughly wiped from wrist to fingertips using a piece of glass wool
soaked in isopropanol. Full details of the sampling methods are given in
Papadopoulou et al. (2016).
After collection, all samples were stored inside a freezer at -20°C until
analysis. The sampling campaign was approved by the Regional Committees for
Medical and Health Research Ethics in Norway (Case number 2013/1269), and
all participants completed a written consent form prior to participation.
3.3 Sample treatment & instrumental analysis
3.3.1 Precautions against contamination
In PE analysis, it is important to avoid contamination in all steps of the
analytical chain (i.e. sampling, storage, extraction, clean up, and instrumental
analysis), which otherwise leads to false high values. For this reason, special
precautions were taken. High-density polyethylene (HDPE) bottles with screw
caps and security lids, pre-cleaned with methanol, were used to store the
samples. Laboratory glassware, sodium sulfate, charcoal and glass wool were
regularly heated at 430 ᴼC overnight and covered with aluminum foil. All
20
organic solvents were tested for PE contamination and no background levels
were detected. Prior to use, SPE cartridges and Teflon tubes were cleaned with
acetone and allowed to dry inside the fume hood. Glass Pasteur pipettes filled
completely with activated charcoal powder and glass wool on the narrow point
were used for filtering the nitrogen (N2) gas stream during evaporation of
organic solvents. Hamilton micro-syringes were used, instead of common plastic
pipettes and tips, for spiking the samples, and aluminum septum vial caps were
used for sample injection.
3.3.2 Urine & nail analysis
Direct analysis of plasticizer metabolites in urine without pre-concentration
was used according to Servaes et al. (2013). All urine spots were creatinine
corrected via a creatinine (urinary) colorimetric assay kit in order to express the
metabolite levels as mg metabolite per g creatinine (mg/gcrea). Prior to extraction
all fingernails were cleaned with acetone in order to remove dust particles and
residues of polish material. Cut pieces of fingernails were then extracted
according a previously published method for nail analysis (Alves et al. 2016a;
Alves et al. 2016c) with a combination of ultrasound assisted extraction and
dispersive liquid-liquid micro-extraction (DLLME).
Urinary and fingernail metabolites were determined using Ultra Performance
Liquid Chromatography (UPLC) coupled to a Waters Xevo TQ-S tandem mass
spectrometer (Waters, Milford, Massachusetts, U.S.) operating in Multiple
Reaction Monitoring (MRM) mode with negative electrospray ionization (ESI-).
Quantification was performed using MassLynx Mass Spectrometry Software
version 4.1 by Waters Corporation. Detailed information on chemical analysis of
urine and fingernails can be found in paper II and its Supporting Information
(SI).
21
3.3.3 In vivo screening
The same urine and finger nail sample extracts were also used to identify
unknown in vivo metabolites of ATBC, DPHP, DEHTP, DEHA, and V6. In
vitro metabolic extracts from each compound of interest were used as reference
solutions to create an in house fingerprint mass spectra database for target
screening in urine and finger nail samples. The in vitro extracts were prepared
individually by incubating each compound with human liver/intestinal
microsomes and/or human liver S9 fractions. Then, the formed metabolites were
extracted by liquid−liquid extraction. Positive and negative controls were also
prepared to monitor the marker activity of key families of enzymes.
For the analysis of in vitro and in vivo extracts, high resolution Q Exactive
Orbitrap LC-MS (Thermo Fisher Scientific, Bremen, Germany) was used for
accurate mass measurements. The ESI source was operated in polarity switching
mode, and product ion scan followed by selective ion fragmentation in data
dependent MS2 (ddMS
2) was used to obtain additional structural information.
Detailed information on chemical analysis of screening emerging pollutants in
urine and fingernails can be found in paper III and its SI.
3.3.4 Air, dust, hand wipe & food analysis
Indoor and personal air samples were extracted according to previous
methods for simultaneous selective detection of PEs and organophosphate esters
(Bergh et al. 2010; Xu et al. 2016). Dust samples were extracted using a
modified method from Bergh et al. (2012) with microwave-assisted extraction
(MAE, Milestone Ethos UP, Sorisole, Italy) under controlled pressure and
temperature program. Dust clean-up was performed using SPE with ENVI-
Florisil SPE cartridges (ISOLUTE FL 500 mg/3 ml from Biotage, Uppsala).
Hand wipes were also extracted by MAE under controlled pressure and
temperature program. After the extraction was completed, all samples were
stored at -20 °C overnight in order to check for and avoid possible lipid
precipitation in the samples after storage. Food analysis for determination of
22
PEs and alternatives was based on an application note from Agilent (Sun 2014).
The homogenized food was extracted with MAE under controlled pressure and
temperature program, and the extracts were cleaned up using suspended
primary/secondary amine phase (removal of polar pigments, sugars and organic
acids) and C18 (removal of lipids and non-polar components).
Samples from different external exposure media (air, dust, hand wipes and
food) were analysed with GC-MS/MS Agilent 7000 (Agilent Technologies,
Santa Clara, California, U.S.). The instrument was equipped with an auto
injector (Agilent 7683B), the injection was in pulsed splitless mode and the
detector was used in Multiple Reaction Monitoring (MRM) mode with electron
impact ionization mode (EI). Quantification was performed with the use of
MassHunter software version B.04.00 for quantitative analysis (Agilent
Technologies, Inc. 2008). Detailed information on chemical analysis of air, dust,
hand wipes and food can be found in paper IV and its SI.
3.4 Quality control
The use of isotopically-labelled standards enabled accurate determination of
target compounds and minimized possible matrix effects. In paper II, mass-
labelled internal standard (IS) solutions for all PE metabolites 13
C-MEHP, 13
C-5-
oxo-MEHP, 13
C-5-OH-MEHP, d4-MiBP, 13
C-MnBP, 13
C-MBzP and 13
C-MEP,
>95% were supplied by Cambridge Isotope Laboratories (Andover, USA).
DINCH metabolites (cx-MINCH and OH-MINCH) and respective deuterated IS
(d2-trans-cx-MINCH and d2-trans-OH-MINCH) were provided by Dr. Koch and
Mr. Schütze (IPA, Ruhr-University Bochum, Germany). In paper IV, the
deuterated internal standards (IS), DMP-d4, DBP-d4, DEHP-d4, and the
volumetric (pre-injection) standard, biphenyl, were bought from Sigma Aldrich
(Steinheim, Germany). Linear calibration curves for all analytes were
constructed each time with correlation coefficients, R2 ≥ 0.99. Blank and
recovery samples were processed together with each batch of samples. The dust
method was evaluated using standard reference material SRM 2585 (household
23
dust) from National Institute of Standards and Technology (NIST, U.S.) as QC
sample.
3.5 Human exposure estimation & risk assessment
The estimated DIinternal of PEs and DINCH were calculated based on the
urinary creatinine corrected concentrations for each metabolite, the 24h urinary
creatinine excretion (CEsmoothed) and the existing 24 h human fractionary
excretion factors (FUE), which were derived from oral doses of the parent
compounds (Table 1). The following two equations were used for the
calculations (Frederiksen et al. 2013; Koch et al. 2003; Mage et al. 2004):
𝐶𝐸𝑠𝑚𝑜𝑜𝑡ℎ𝑒𝑑(𝑔 𝑑𝑎𝑦⁄ ) = A × [(140 − 𝑎𝑔𝑒(𝑦𝑒𝑎𝑟)] × 𝐵𝑊(𝑘𝑔)1.5 ×
ℎ𝑒𝑖𝑔ℎ𝑡(𝑐𝑚)0.5 × 10−6 (1)
𝐷𝐼𝑖𝑛𝑡𝑒𝑟𝑛𝑎𝑙 (𝜇𝑔
𝑘𝑔𝐵𝑊×𝑑𝑎𝑦) =
𝛴[𝑈𝐸𝑚𝑒𝑡𝑎𝑏𝑜𝑙𝑖𝑡𝑒 𝑐𝑟𝑒𝑎(𝜇𝑔 𝑔𝑐𝑟𝑒𝑎)⁄
𝑀𝑊𝑚𝑒𝑡𝑎𝑏𝑜𝑙𝑖𝑡𝑒(𝜇𝑔 µ𝑚𝑜𝑙⁄ )]×𝑀𝑊𝑃ℎ𝑡ℎ𝑎𝑙𝑎𝑡𝑒(𝜇𝑔 𝜇𝑚𝑜𝑙⁄ )×𝐶𝐸𝑠𝑚𝑜𝑜𝑡ℎ𝑒𝑑(𝑔 𝑑𝑎𝑦⁄ )
𝛴𝐹𝑈𝐸×𝐵𝑊(𝑘𝑔) (2)
where in Equation 1, 140 and A (1.93 for males and 1.64 for females) are
constants simplifying for the body surface area; the individual height (cm), body
weight (kg) and age for each participant, while in Equation 2, UE is the
concentration of the metabolite in urine (creatinine adjusted); MWPhthalate is the
molar mass of the parent diester; MWmetabolite is the molar mass of the
corresponding metabolite.
The total daily intake based on external media DIexternal [ng/kg/d] was
calculated as the sum of intake through air inhalation (both gaseous and
particulate phase), dust and dietary ingestion and dermal absorption (sum of
dermal uptake from air, dust and PCPs):
𝐷𝐼𝑒𝑥𝑡𝑒𝑟𝑛𝑎𝑙 = 𝐷𝐼𝑖𝑛ℎ + 𝐷𝐼𝑑𝑢𝑠𝑡 + 𝐷𝐼𝑑𝑖𝑒𝑡 + 𝐷𝐼𝑑𝑒𝑟𝑚𝑎𝑙 (3)
24
Table 2. Established health limit values for phthalates and DINCH (TDI,
tolerable daily intake; RfD, reference dose).
Compound
Health limit values
[µg/kg/day] Reference
TDI RfD
DMP - 375* (ECHA 2012; Gray et al. 2000)
DEP - 800 (Brown et al. 1978; USEPA 1987c)
DiBP 10**
200 (Kortenkamp and Faust 2010)
DnBP 10 100 (EFSA 2005b; USEPA 1987b)
BBzP 500 200 (EFSA 2005c; USEPA 1989)
DEHP 50 20 (EFSA 2005a; USEPA 1987a)
DINP 150 - (EFSA 2005d)
DIDP 150 - (EFSA 2005e)
DPHP 200 - (UBA 2015)
DINCH 1000 700 (Bhat et al. 2014; EFSA 2006)
*Calculated
**By analogy to DnBP
To assess the risk for humans, we calculated the hazard quotient (HQ)
according to Equation 4, that is, the ratio between the total daily intake
estimated via external exposure media or biomonitoring data, and health limit
values, expressed as oral reference doses (RfDs) or tolerable daily intakes
(TDIs) established by EFSA or USEPA (Table 2).
𝐻𝑄 =𝐷𝐼
𝑇𝐷𝐼 𝑜𝑟 𝑅𝑓𝐷 (4)
Moreover, cumulative risk assessment from combined chemical exposure
25
doses was performed by using the hazard index (HI), which is expressed by:
𝐻𝐼 = ∑ 𝐻𝑄𝑖𝑛𝑖=1 (5)
where n is the number of substances, and values below 1 are consider safe
(Kortenkamp and Faust 2010). Only the potential anti-androgenic chemicals,
DiBP, DnBP, BBzP and DEHP, were included in the HI calculation.
3.6 Potential predictors of exposure
Information on sociodemographic and lifestyle characteristics that might
affect the concentrations of PE and DINCH metabolites in adults was collected
by the A-TEAM questionnaires. Different categories were made such as gender,
age, educational level, smoking status (yes/no), coloring hair (yes/no), average
weekly frequency of washing hair, average daily frequency of washing hands
and eating with hands (e.g. eating sandwiches or other snacks without using
cutlery), using gloves when cleaning the house (yes/no), and living in a house
with PVC in walls/floors (yes/no). In addition, the total number of individual
PCPs and total number of times PCPs were applied to the hands during the last
24 h were categorized into two groups (≤ 5 PCPs vs > 5 PCPs and ˂ 5 times/day
vs ≥ 5 times/day). The total number of foods wrapped/enclosed in plastic
packaging during the two consecutive sampling days (<10 foods/day vs ≥10
foods/day) was also categorized.
To explore the associations between all these different categories, and
concentrations of individual PE and DINCH metabolites in urine and
fingernails, we applied multivariate linear mixed models with a random
intercept per participant and a fixed effect on all other characteristics. The
relative change (and 95% CI) of individual phthalate metabolites per unit
increase or per category of the different characteristics was calculated as: (exp
β-1) ×100, where β is the linear regression coefficient. Correlations, differences
and associations were considered significant if p < 0.05 (for the identification of
potential determinants of phthalate metabolite concentrations, p < 0.2 were also
26
reported). All statistical analyses were performed using STATA 14 (StataCorp,
College Station, TX) and SPSS (SPSS for Windows, version 22.0, SPSS,
Chicago, IL).
27
4 Results & discussion
4.1 Alternative plasticizers to phthalate esters
Almost all of the alternative plasticizers reviewed in paper I are listed as
high production volume (HPV) chemicals in the 2007 OECD HPV list
(produced or imported at volumes ≥ 1000 tonnes/year in at least one EU
country/region) or in the chemicals in the U.S. HPV challenge program list.
Moreover, some alternatives are being produced in very large quantities in the
EU for use in multiple applications, such as DINCH (10 000+ tonnes/year) or
DEHT (10 000–100 000 tonnes/year). In Sweden, there is an increasing trend in
the total use of alternative plasticizers compared to common PEs over the last 10
years, especially for DINCH and DPHP. As shown in Figure 2, the total use of
PEs remained fairly constant until a sudden decrease in 2010, when the use of
DPHP rapidly increased, which illustrates the use of DPHP as a substitute for
commonly used PEs. The total use of alternative plasticizers was about 3 times
lower than PEs in 1999 and remained at the same level until a substantial
increase from 2011 to 2012, surpassing the total use of PEs, which was mainly
attributable to the increased use of DINCH (Figure 4, paper I). Hence, there has
been a shift from using PEs to alternative plasticizers, and alternatives will most
likely continue to increase in the coming years due to PE restrictions and phase
outs.
28
Figure 2. Total use of commonly used phthalate esters (sum of DEP, DnBP,
DiBP, BBzP, DEHP, DINP and DIDP) and the more recently occurring DPHP
compared to the use of alternative plasticizers (sum of substances listed in Table
2 of paper I excluding COMGHA) in chemical products in Sweden from 1999
to 2012 (figure reproduced from paper I with permission from Elsevier).
Physicochemical properties of alternative plasticizers vary considerably
depending on the substance group or the individual chemical. Many of them will
likely partition to organic carbon rich matrices (soil, biota, dust etc.) because of
high log KOW and log KOA values. Similarities between common PEs and
alternatives exist for some cases (Figure 2, paper I).
When assessing sources, fate and exposure to plasticizers, the indoor
environment is important since these chemicals are mostly found in consumer
products used in homes. Indoor fate was assessed using the Stockholm
Multimedia Urban Fate (SMURF) model (Cousins 2012). According to the
model predictions (Figure 5, paper I), alternative plasticizers are more likely to
be found in dust on vertical (e.g. walls and ceilings) and horizontal (e.g. floors)
surfaces than in indoor air, similar to the selected PEs (i.e. DEHP, DiNP, DiDP,
DPHP). Floor dust is commonly removed by vacuuming and mopping on a more
29
regular basis than cleaning of walls and ceilings. Hence, chemicals attached to
horizontal surfaces have the potential to be removed quicker and in higher
amounts than those sorbed to vertical surfaces. Based on these results, exposure
to most PEs and their alternatives is more likely to occur via dermal uptake or
dust ingestion as these substances will not partition strongly to indoor air, except
trimethyl pentanyl diisobutyrate (TXIB) and glycerin triacetate (GTA), where
exposure via inhalation will be more important. This indoor fate modeling did
not provide information for dietary intake which is usually the most important
route of human exposure. Generally, alternative plasticizers can be expected in
various exposure matrices in the indoor environment, although only very few
exhibit strong partitioning to indoor air. Instead, it is likely that many
alternatives will be detected in dust and/or food due to their low volatility and
relatively high hydrophobicity.
Regarding the human exposure to alternative plasticizers, data are lacking
and clear evidence for exposure only exists for DEHT and DINCH, which show
increasing trends in body burdens (Nagorka et al. 2011; Schutze et al. 2014).
Human intake rates were collected and compared with established threshold
values resulting in risk ratios below 1, indicating low risk (Tables 3-6, paper I),
except for infant's exposure to epoxidized soybean oil (ESBO). More attention
should be given to DINCH because only the dietary uptake was consider for this
assessment (NICNAS 2011) while other important routes of exposure were not
included (e.g. transdermal exposure and dust ingestion), and total exposure to
DINCH may be close to the TDI or RfD.
Finally, PBT properties of the alternatives indicate mostly no reasons for
concern. None of the selected alternative plasticizers displayed a PBT or very
persistent, very bioaccumulative (vPvB) profile according to the REACH
criteria (Table 7, paper I) which indicates no reason for concern, apart from
tris(2-ethylhexyl) phosphate (TEHPA) which was estimated to be persistent and
tricresyl phosphate (TCP) to be toxic.
30
4.2 Urine & nails
Most of the urinary metabolites were frequently detected in all urine spots,
except the hydrolytic monoesters of DEHP, DPHP and DINCH, where the
oxidative metabolites were abundant. The urinary levels for all PEs in the
Norwegian cohort were comparable, but in most cases lower than in other
biomonitoring studies around Europe, North America and Australia (Table SI3,
paper II). The determined levels of DINCH oxidative metabolites in this study
population, together with recently obtained data from children and adults in U.S.
and Germany (Fromme et al. 2016; Schutze et al. 2014; Silva et al. 2013a)
indicate that the increasing production volumes of DINCH implies increased use
in products, which has a direct impact on human exposure.
In contrast to urine, the most abundant metabolites in nails were the
hydrolytic PE monoesters, while the presence of DEHP and DINCH oxidative
metabolites was low (Table 1, paper II), which was in agreement with other
studies (Alves et al. 2016a; Alves et al. 2016b; Alves et al. 2016c). These results
indicated that there might be equilibrium between accumulation of the same
metabolite in nails versus its excretion via urine. However, this can be
dependent on several intrinsic factors (e.g. age, gender, etc.), as well as the
diffusion from blood to nails (metabolic rate and transfer rate). The fast
excretion of oxidative metabolites in all three urine spots and the frequent
presence of hydrolytic monoester metabolites in nails, suggested that
bioaccumulation versus excretion were accountable for different proportions
depending on the metabolites formed (hydrolytic or oxidative). To our
knowledge, it is still not clear whether PEs are first transported to the nails and
then metabolized or metabolized somewhere else in other body reservoirs (e.g.
blood) and then transported to the nails. Therefore in order to understand how
the PE metabolites are partially accumulated in nails, future studies on
metabolism and exposure assessment in nails (and urine) are needed.
Regarding the correlations between urine and nails, only MEP (specific
metabolite of DEP) was highly and consistently correlated probably due to
higher and constant human exposure to consumer products containing DEP as it
31
is not regulated as other PEs. For the other PEs and DINCH, we observed no
correlations, which is another indication that further investigation is warranted
in terms of mechanisms of transfer and/or incorporation pathways of plasticizers
into nails.
4.3 Potential predictors of exposure
Findings in paper II suggested that participants who smoked had higher
levels of some PE metabolites in urine and nails than non-smokers. Although
this was the first time that this association has been reported, the limited number
of smokers (n=4) did not allow us to draw strong conclusions. Further
investigation is required to study the significance of this finding since other co-
factors, such as the hand-to-mouth contact during smoking or contact with the
material of cigarette filters might contribute to and enhance this association.
Figure 3. Median urinary MEP concentrations (in μg/g creatinine) in three
consecutive urine spots related to the number of personal care products used in
32
the last 24 hours (figure reproduced from paper II with permission from
Elsevier).
The more frequent use of PCPs was associated with an increase of
monomethyl phthalate (MEP) in urine (Figure 3), while washing the hands more
frequently reduced levels of DEHP and DINCH oxidative metabolites. Eating
with hands, and wearing plastic gloves when cleaning the house were positively
associated with elevated PE metabolite levels in urine and nails. Also, each year
increase in age was associated with greater PE exposure. Participants who
consumed ≥ 10 foods items/day stored in plastic packaging during the two
sampling days, had elevated levels of DEHP metabolites in urine, while in nails
this association was the inverse. Differences between urine and nails might be
explained by the different exposure windows they reflect, by differences in
metabolic and kinetic behavior of PE biomarkers in the different matrices, but
also by possible external contamination influences on the PE biomarkers.
4.4 Air, dust, hand wipes & food
PE and DINCH concentrations were determined in different external
exposure media (Table 2, paper IV). High molecular weight PEs and DINCH,
which are used in floor and wall coverings or in upholstered furniture, were
found in house dust in high concentrations. Low molecular weight PEs were
highly detected in indoor and personal air samples. In hand wipes, DEHP and
DiNP were the predominant plasticizers, while DEP was present in all samples
probably due to the use of PCPs. DPHP and DINCH, the main substitutes of
DEHP in food processing equipment and package materials, were present in
food with high detection frequencies (DF %) similar to all PEs.
33
4.5 Human exposure estimation & risk assessment
Figure 4. Daily intakes [ng/kg/d] calculated from internal exposure based on
biomonitoring data (red) and from external exposure pathways (blue). Box plots
represent 25th
, 50th
(shown by the horizontal line) and 75th
percentiles. Whiskers
indicate maximum and minimum values within 1.5 times the inter-quartile range
and circles show outliers. A * indicates a significant difference between the two
datasets in a Mann-Whitney-U test (p < 0.05).
Compared to the internal exposure assessment using urinary metabolites
(paper II), the external exposure estimation (paper IV) resulted in higher or
equal DIs (Figure 4). This means we successfully linked concentrations from
external matrices to internal concentrations. In addition, exposure to PEs and
DINCH is well understood for adults as even significant differences between the
two approaches were not large. Hence, we concluded that we had captured the
most relevant uptake pathways with the applied external exposure assessment
method. In general, all estimated DIs to PEs and DINCH were below the
established health limit values (Figure 4). The internal exposure to PEs was
lower than other European studies (paper II). It is difficult to compare the
34
results of external exposure (paper IV) to other international studies because
most of them focused on either one or a few exposure pathways and did not
compare all potentially relevant routes.
Risk assessments performed for the Norwegian study population based on
internal (paper II) and external (paper IV) exposures were in good agreement.
All the participants had HQs with median values well below 1 for all
plasticizers, meaning low risk (Figure 5). The highest HQ was 0.65 regarding
the internal exposure to DnBP, which was still below the value of 1. Moreover,
DPHP and DINCH, the two recent alternative plasticizers with an expected
increase in exposure in the coming years, showed low risk.
The cumulative risk assessments via internal exposure (HI median of 0.11
based on TDI, and 0.04 based on RfD) and via external exposure (HI median of
0.2 based on TDI, and 0.08 based on RfD) indicated a low anti-androgenic risk
potential (Figure 5), even for the worst case scenario by using the maximum DI
estimates. Even though HI < 1 means no risk, there are more substances with
similar modes of action/end-points which were not included in these cumulative
risk assessments. PEs considered are not the only chemicals with potential anti-
androgenic or endocrine disruption properties. Thus, a more accurate HI
estimation should take into account the additional effects of other suspected
chemicals, such as bisphenol A, parabens, triclosan, polychlorinated biphenyls,
which exhibit similar adverse health effects (Kortenkamp and Faust 2010;
Maffini et al. 2006; Schug et al. 2011).
35
Figure 5. Hazard quotients (HQs) based on established health limit values
(TDIs and RfDs). Hazard indexes (HIs) of plasticizers (i.e. DiBP, DnBP, BBzP
and DEHP) with similar adverse health effects. Box plots represent 25th
, 50th
(shown by the horizontal line) and 75th
percentiles. Whiskers indicate 5th
and
95th
percentiles.
36
4.6 Relative importance of different exposure pathways
The relative importance of dermal uptake, dust ingestion, inhalation and
dietary intake was elucidated for PEs and DINCH (paper IV, figure 2). In all
cases, diet was the predominant pathway with a contribution range of 55-95% to
total intake. Especially for high molecular weight PEs and DINCH dietary
intake contributes more to total intake. Inhalation contributes approximately 15-
45% of low molecular weight PEs. Unsurprisingly, this uptake pathway is
negligible for higher molecular weight PEs, since they are less volatile than
lower molecular weight compounds. Dust ingestion on the other hand was more
important for higher molecular weight compounds as expected (approximately
40% contribution for DINP intake; 10% for DEHP, DPHP and DINCH intake;
and 5% for BBzP intake). Total dermal uptake (from air, dust, PCPs) was very
small in all cases and did not substantially influence total intake, except for DEP
where the dermal uptake contributes to approximately 10% of total intake,
because DEP is a common additive in PCPs (Guo and Kannan 2013; Koniecki et
al. 2011).
Dermal uptake was assessed using hand wipe samples which might have
underestimated real exposure as hand wipe concentrations are not always
constant and chemicals already absorbed by the skin could not be captured. It is
possible that dermal penetration had already occurred for the majority of
chemical masses on the skin and hand wipes only represented the unabsorbed or
slowly absorbed fraction. It was clear that air and dust were not important
sources for dermal exposure, while frequent use of PCPs can be regarded as a
major contributor (paper IV, figure 3 & Table SI2). Overall, dermal exposure is
still poorly understood and therefore has much higher uncertainty than other
estimations.
4.7 In vivo screening
After in vitro metabolism of four alternative plasticizers (i.e. DPHP,
DEHTP, ATBC, and DEHA) and one flame retardant (i.e. V6) using human
37
liver/intestinal microsomes and/or human liver S9 fractions, a set of 13
metabolites was selected for further identification in urine and finger nails using
high-resolution Q Exactive Orbitrap LC-MS (in vitro extract metabolites are
shown in Table SI-2, paper III). The identified analytes and their DF% in finger
nail and urine samples are presented together with the interpretation of their
fragments in paper III.
For DPHP, no metabolites corresponding to those generated from in vitro
experiments with DPHP were detected in finger nail or urine samples. In
general, DPHP oxidative metabolites are excreted rapidly in urine, similar to
DEHP oxidative metabolites (Preuss et al. 2005). Therefore, finger nails do not
seem to be a suitable matrix for DPHP exposure measurements and sensitive
(extraction/analytical) tools are needed for measuring DPHP oxidative
metabolites in urine.
For DEHTP, two metabolites were identified in nails and/or in urine, namely
mono(2-ethylhexyl) terephthalate (MEHTP) and 1-mono-(2-ethyl-hydroxy-
hexyl) benzene-1,4-dicarboxylate (OH-MEHTP). Several oxidative DETHP
metabolites were identified as promising biomarkers in humans in vitro (human
liver microsomes) or in vivo (urine) in previous studies (Lessmann et al. 2016;
Silva et al. 2015), but this was the first time that DEHTP metabolites have been
explored and detected in finger nails. Therefore, in the case of DEHTP, results
in paper III suggested that finger nails could be a possible alternative matrix for
biomonitoring.
ATBC was detected only in nail samples, but it was difficult to determine
whether ATBC levels found in nails were coming from internal or external
exposure. Also, three metabolites of ATBC and their isomers were identified in
nails. The complete picture for ATBC exposure was unclear, and because there
were no individual standards for each of the identified metabolites, it was
difficult to elucidate which specific metabolite is rapidly excreted in urine
and/or accumulated in nails. Overall, these results indicate that nails may be a
suitable matrix for non-invasive monitoring of ATBC exposure, preferably by
38
measuring its hydrolytic and oxidative metabolites, but further investigation is
required.
To our knowledge, paper III was the first study where metabolites of
DEHA have been both generated in vitro and further detected in urine and nails
with explanation of their fragments. Previously, DEHA metabolites were
measured in vivo in urine or breast milk (Cousins et al. 2007; Fromme et al.
2007; Fromme et al. 2013). Also, one hydrolytic and three oxidative metabolites
of DEHA were identified in vitro with human liver microsomes and then further
detected in urine samples (n = 144) from American adults (Silva et al. 2013b).
In paper III, the hydrolytic metabolite of DEHA (MEHA) was mainly
identified in finger nails, while the oxidative metabolite of DEHA (MEHHA)
was shown to be mostly present in urine. Hence, results from in vitro
experiments were able to partly predict the DEHA metabolism in humans.
Although V6 could be identified in finger nails with low DF %, there were
limitations of the study (i.e. same extraction methods were applied as for
plasticizers without method evaluation). This may reflect the low
bioaccumulation potential of V6 in nails. However, several issues need to be
addressed in future research to enhance the understanding of human exposure to
V6 and the use of nails in HBM studies.
39
5 Concluding remarks
In paper I, according to the reviewed knowledge, the use of alternative
plasticizers seems to be of low risk for humans as calculated risk ratios were
mostly well below 1. None of the important alternative plasticizers included in
analysis of A-TEAM samples displayed a PBT or vPvB profile according to the
REACH criteria (ECHA 2014). However, their constant production and
emission can also lead to continuous exposure and pseudo-persistence, similar to
PEs. Furthermore, results in paper I indicated that the majority of the
alternative plasticizers are likely to partition to organic matter on surfaces and
airborne particles indoors. Settled dust and/or food can be important sources of
human exposure due to hydrophobic characteristics of these chemicals. DINCH
levels measured in house dust and food (paper IV), but also other alternative
plasticizers measured in several studies (Fromme et al. 2013; Fromme et al.
2016; Larsson et al. 2017) further confirmed this finding.
In papers II & IV, low exposure to PEs and DINCH was observed for the
included Norwegian population, based on estimated daily intakes using
concentrations of their metabolites in urine. The cumulative risk assessment for
combined plasticizer exposure, expressed as HQs and HI, was below established
risk limits, even considering the worst case scenario (i.e. maximum values).
However, we should not ignore that there are several other chemicals with
similar adverse health effects that could contribute to this cumulative risk
assessment. The determined concentrations of PE metabolites in urine were in
general lower than reported in other European studies (Dewalque et al. 2014;
Myridakis et al. 2015; Saravanabhavan et al. 2013; Wittassek et al. 2007b),
while concentrations of DINCH metabolites were comparable to reported levels
from 2012, and higher than those reported before 2012 (Fromme et al. 2016;
Gomez Ramos et al. 2016; Schutze et al. 2014; Silva et al. 2013a). The
decreasing PE exposure due to regulatory restrictions and market changes (Goen
et al. 2011; Zota et al. 2014) was confirmed by the findings in papers II & IV,
40
while exposure to alternative plasticizers, such as DPHP and DINCH is
expected to increase in the near future (Schutze et al. 2014).
Age, smoking, use of PCPs and many other everyday habits, such as
washing hands or eating food from plastic packages were identified as possible
contributors to plasticizer exposure, but still no firm conclusions could be drawn
due to the small sample size of the study population (paper II).
Finger nails may be a useful non-invasive matrix for human biomonitoring
of specific organic contaminants, but further validation is needed in terms of
mechanisms of transfer and/or incorporation pathways of plasticizers into nails
(paper II and III). Measurements of plasticizer metabolite levels in nails can
only be regarded as a first step in establishing this matrix as a future matrix for
exposure assessment and risk characterization. This matrix could extend the
window of exposure measurement for short lived chemicals that are rapidly
excreted in urine.
Paper III suggested that in vitro assays can reliably mimic in vivo processes
since metabolites from alternative plasticizers and one flame retardant present in
in vitro extracts were further identified in vivo in urine and finger nail samples.
Paper IV revealed that external exposure media such as air, dust, food and
hand wipes still have significant concentrations of PEs despite phase-outs or
restrictions for most of these chemicals. Human body burden is not negligible,
and external and internal exposures were successfully linked and found to be in
reasonable agreement. Dietary intake seemed to be the most important exposure
pathway for the Norwegian study population, followed by inhalation and dust
ingestion. Overall, the human risk was low, which is in agreement with the
biomonitoring findings in paper II.
41
6 Future perspectives
In contrast to commonly used PEs, important physicochemical properties of
new alternative plasticizers, such as vapor pressure, solubility in water or log
KOW, have not been extensively measured. The selection of a property value, in
order to model the indoor fate of these chemicals, can be difficult since not
many studies exist for comparison. Also, EPISuite estimations sometimes vary
by more than two orders of magnitude from real experimental results. Therefore,
additional experimental measurements of these key properties are necessary for
a more reliable indoor fate assessment of alternative plasticizers.
To our knowledge, there is no other human cohort analyzed to such an extent
as the A-TEAM cohort. Similar comprehensive datasets with biomonitoring data
and levels from different external exposure media are needed to link the external
and internal exposure to plasticizers for vulnerable sub-group populations, such
as children and pregnant women, since they have different physiological and
exposure factors than adults. Such comprehensive results could then be used
together with scientifically based tools and approaches to support public health
decisions and policy making.
Certain routes of human exposure are still poorly understood because they
are not explicitly modeled. These include the dermal exposure via direct contact
with plastic materials, and hand-to-mouth contact, where dust ingestion may
also contribute, apart from the skin concentration which can be measured
directly with hand wipes. Overall, dermal exposure requires further
investigation. There is a need for additional skin permeation experiments for
some plasticizers, such as DEP, BBzP and the alternatives DPHP and DINCH,
in order to confirm their skin metabolization. Then, more accurate skin
absorption fractions could be generated and used in future dermal exposure
assessments as current values are likely quite low. Likewise, dietary uptake
estimates will improve in accuracy, when in vitro results from bioaccessibility
experiments regarding the gastrointestinal absorption of PEs clarify
uncertainties, which currently potentially lead to overestimation.
42
Alternative plasticizers have been recently introduced to the market to
replace critical PEs under restriction. Hence, an increasing trend in
concentration in humans is expected, without knowing when peak exposure will
be reached. Careful and continuous monitoring of important alternative
plasticizers, such as DPHP, DINCH, DEHTP, ATBC and DEHA, are required to
assess body burdens of these emerging contaminants in the human population,
so that rapid control and prevention can be achieved to minimize the risk.
Furthermore, special consideration should be given to sources such as dust or
children's toys and more current studies that represent high risk groups (e.g.
children) are required. This is an important aspect especially for DINCH, which
is used as an alternative plasticizer in children's toys.
Finger nail analysis indicated that most of the targeted plasticizers can
accumulate in this matrix. However, it is still uncertain whether the levels
measured in finger nails reflect internal and/or external exposure since
contributions are difficult to distinguish. Bioaccessibility experiments should be
conducted in order to check possible hydrolysis which might occur directly on
the nail plate. Further understanding is required about metabolism in nail plate
and/or surrounding tissues (e.g. nail bed), and the excretion versus the
bioaccumulation rate, which might be chemical specific. Toe nails would also be
relevant to analyze, as the levels and exposure are usually distinct. Finger nails
are more exposed to external sources than toe nails, which may hamper the
identification and understanding of internal versus external exposure. Replicate
nail analysis over a long period of time, would provide a better understanding
about the stability and reliability of the levels reported in different nail
segments, and how well nails can be considered good biomarkers of past
exposure.
In urine, the metabolites of DPHP, DINCH and DEHTP were only recently
identified, and their excretion rates derived after single oral dose experiments
(Koch et al. 2013; Leng et al. 2014; Lessmann et al. 2016). In the same way,
experiments should be done to identify specific biomarkers of exposure for other
important alternative plasticizers, such as ATBC and DEHA. On top of that,
43
further investigation is also required on the toxicological effects of all these new
metabolites. To ensure proper quantification of exposure in future biomonitoring
studies, the development and validation of new analytical methods and
commercially available reference standards for these specific metabolites are
essential.
44
7 Acknowledgments
The research leading to this thesis has received funding from the European
Union Seventh Framework Programme FP7/2007–2013 under Grant agreement
no 316665 (A-TEAM project).
The greatest thank you goes to my advisors Dr. Jörgen Magnér at IVL
Swedish Environmental Research Institute and Prof. Cynthia A. de Wit at the
Department of Environmental Science and Analytical Chemistry (ACES),
Stockholm University, who supported my work. I gratefully appreciate that you
trusted me with such a variety of different tasks within my PhD project, and let
me work independently. This way you gave me a boost of self-esteem and
confidence which are important qualifications for my future professional career.
Thank you to all my IVL colleagues (especially Anna Palm-Cousins,
Momina Bibi, Mikael Remberger and Lennart Kaj) for the first three years of
my PhD. I am very grateful for having the privilege of working with you, extend
my knowledge and learn new competences.
Thank you to Stefan Voorspoels and my A-TEAM colleague Andreia Alves
for all their orientation during my secondment at Flemish Institute for
Technological Research (VITO). Also, I gratefully acknowledge Prof. Adrian
Covaci at Antwerp University (UA) for his advices and fast feedback.
The credit also goes to Thuy Bui, Katerina, Fenix and rest of my A-TEAM
colleagues Stathis, Melissa, LuRa, Ana, Somrutai, Joo Hui, Lila, Juan, Claudio
and Gopal. I am very grateful for having the opportunity to discuss with you
about science, exchange knowledge and make collaborations. Thank you for the
moments we shared during and after the A-TEAM meetings.
I would like to say thank you to all the people at ACES for providing a nice
working atmosphere during the last year of my PhD. In particular, I would like
to thank Ulrika Nilsson as I had the pleasure to collaborate with you and I am
very grateful for your contribution in paper III.
Lastly, I would like to thank my family and friends for their support over the
years. Special thanks to Eleni M. for unconditional support and understanding.
45
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