Calibration and validation of a microporous polyethylene passive sampler for quantitative estimation of illicit drug and pharmaceutical and personal care product (PPCP) concentrations in wastewater influent

Journal Pre-proof Calibration and validation of a microporous polyethylene passive sampler for quantitative estimation of illicit drug and pharmaceutical and personal care product (PPCP) concentrations in wastewater influent

Sarah McKay, Ben Tscharke, Darryl Hawker, Kristie Thompson, Jake O'Brien, Jochen F. Mueller, Sarit Kaserzon PII:

S0048-9697(19)35886-3

DOI:

https://doi.org/10.1016/j.scitotenv.2019.135891

Reference:

STOTEN 135891

To appear in:

Science of the Total Environment

Received date:

4 August 2019

Revised date:

27 November 2019

Accepted date:

30 November 2019

Please cite this article as: S. McKay, B. Tscharke, D. Hawker, et al., Calibration and validation of a microporous polyethylene passive sampler for quantitative estimation of illicit drug and pharmaceutical and personal care product (PPCP) concentrations in wastewater influent, Science of the Total Environment (2019), https://doi.org/10.1016/ j.scitotenv.2019.135891

This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

© 2019 Published by Elsevier.

Journal Pre-proof

Calibration and validation of a microporous polyethylene passive sampler for quantitative estimation of illicit drug and pharmaceutical and personal care product (PPCP) concentrations in wastewater influent Sarah McKaya, Ben Tscharkea, Darryl Hawkera,b, Kristie Thompsona, Jake O’Briena, Jochen F. Muellera

a

of

and Sarit Kaserzona The University of Queensland, Queensland Alliance for Environmental Health Sciences (QAEHS), 20 Cornwall

Griffith University, School of Environment and Science, 170 Kessels Road, Nathan, QLD, 4111 Australia

re

-p

b

ro

Street Woolloongabba QLD 4102, Australia

Keywords

lP

Drug monitoring, passive sampling, wastewater analysis, sewer epidemiology, sampling rates

na

Highlights

Microporous polyethylene passive samplers deployed in WWTP influent for ≤ 29 days



Uptake of illicit drugs and PPCPs allowed derivation of sampling rates



Drug/PPCP levels in separate deployment predicted well with derived sampling rates



Sampler shows promise for providing necessary drug monitoring data in wastewater

1

Jo

ur



Journal Pre-proof

Abstract Wastewater-based epidemiology (WBE), the per capita normalised measurement of drugs, chemicals or metabolites in wastewater influent, relies on sampling and quantitative analysis to evaluate temporal and spatial trends of chemical consumption. Continuous, high-resolution, flow proportional composite sampling is optimal for accurate representations of chemical mass loads, but is rarely implemented, with conventional autosamplers providing relatively low frequency time or volume proportional samples. However, due to equipment or resource constraints at many

of

wastewater treatment plants (WWTPs), even this may not be feasible. Passive sampling may provide an alternative sampling strategy. To investigate this, samplers comprising hollow, cylindrical

ro

Microporous Polyethylene Tubes (MPTs) containing polymeric sorbent phases of Strata-X and Strata-

-p

X in agarose were simultaneously deployed in a municipal WWTP influent stream. Samplers were extracted, analysed and evaluated for a range of illicit drugs and PPCPs after 4, 7, 15, 21, and 29 day

re

deployments. The MPT samplers were calibrated against 24-hour time proportional composite grab samples that were collected in parallel. Diffusion through the MPT governed uptake, reducing or

lP

eliminating the influence of external flow rates that may fluctuate unpredictably in a WWTP environment. Calibration data for six illicit drugs and fourteen pharmaceuticals and personal care

na

products (PPCPs) demonstrated linear accumulation in the samplers (R2 ≥ 0.84), including methamphetamine, benzoylecgonine, MDMA, codeine and carbamazepine. Derived sampling rates

ur

for the analytes ranged from 0.25 to 17 mL d-1 for ibuprofen and verapamil, respectively. A validation study using this sampling rate data showed the MPT could effectively quantify

Jo

concentrations (≥ 0.1 ng mL-1) of a range of amphetamine-type stimulants, opioids and metabolites as well as nicotine, accounting for 95% of the variance in parallel composite grab sample concentrations of these compounds. The MPT sampler shows promise for providing essential monitoring data for WBE, informing future intervention and research strategies.

1. Introduction Wastewater-based epidemiology (WBE) has become an accepted method for surveying population consumption/exposure of a broad suite of chemicals. Of particular interest is its ability to identify community consumption of illicit drugs (Zuccato et al., 2005; Choi et al., 2018). The premise of WBE is that when humans consume chemicals either intentionally or unintentionally, these compounds are excreted from the body into wastewater either as the parent compound or 2

Journal Pre-proof metabolites. By sampling the wastewater at the inlet of municipal wastewater treatment plants (WWTPs) and measuring the concentrations of these chemicals along with measurements of daily flow volumes and the population size, chemical excretion factors can be applied and ultimately estimates of per capita mass load of a chemical consumed by the WWTP catchment population can be achieved (Zuccato et al., 2005). One of the principal uncertainties in wastewater based monitoring programs concerns the representativeness of the collected samples (Castiglioni et al., 2013). Sample collection should preferably represent and integrate the sampling period and the concentration variance of chemicals (typically over 24 hours or a week). Passive samplers are designed to accumulate target chemicals

of

over an extended period, providing time-weighted-average water concentration estimates (Booij et

ro

al., 2007; Kaserzon et al., 2014). They typically work via a chemical diffusion gradient, resulting in the movement of chemicals into the sampler’s receiving phase (Alvarez et al., 2004; Huckins et al., 2006;

-p

Kingston et al., 2000). Passive samplers have been calibrated extensively for many chemicals and environments but few have focused on wastewater in WWTPs that poses particular challenges. The

re

Polar Organic Chemical Integrative Sampler (POCIS) has been a predominant tool in those passive

lP

sampling investigations that have been conducted in municipal wastewater influent to date, (Harman et al., 2012; Harman et al., 2011). POCISs have been used for the analysis of more than 300 polar organic compounds in water including illicit drugs, pesticides and pharmaceuticals (Gonzalez-

na

Rey et al., 2015; Harman et al., 2012; Jones-Lepp et al., 2004; Metcalfe et al., 2011; Morin et al., 2012). From the WBE perspective they have also been used for the long-term monitoring of

ur

community level illicit and prescription drug consumption (Baz-Lomba et al., 2017).

Jo

However, there remain some aspects of such sampling devices that require further development and optimisation (Fauvelle et al., 2017b; Harman et al., 2012). For example, fluctuating water velocity can substantially alter the thickness of the water boundary layer (WBL) at the surface of the POCIS and similar samplers, compared to that of the membrane (Fauvelle et al., 2017a; Fauvelle et al., 2017b). At relatively high linear velocities over the sampler surface e.g. > 2.0 cm s-1 with POCIS, the WBL thickness approaches a lower bound and sampling rates are not significantly affected (Di Carro et al., 2014). If the water velocity is not sufficiently great however, and the rate of transfer of chemicals into the sampler is influenced by diffusive transit of the WBL, water velocity fluctuations can affect the analyte sampling rate values used to derive estimates of water concentrations (Booij et al., 2007; Fauvelle et al., 2017a; Fauvelle et al., 2017b; Vrana et al., 2006). This is a consideration for WWTP deployments where influent is prone to water velocity fluctuations (Salgado et al., 2011). The accumulation of analytes in the membrane rather than the sorbent phase of the POCIS is 3

Journal Pre-proof another feature that is often overlooked (Harman et al., 2012). Wastewater also typically contains relatively high levels of microorganisms, dissolved and suspended solids as well the possibility of biofilm formation. All these factors may compromise passive sampler performance. An approach to address the influence of fluctuating water flow is to create a higher resistance to mass transfer of chemicals passing through the sampler itself. The relative importance of diffusion through the WBL should then be decreased, providing more stable sampling rates and reduced uncertainty in water concentration estimates. In response to these issues, we developed a passive sampler for glyphosate and its primary metabolite aminomethyl phosphonic acid using a Microporous Polyethylene Tube (MPT) (Fauvelle et al., 2017b). The MPT minimises the effect of

of

external flow rate and the WBL on analyte uptake by increasing the diffusive resistance to chemicals

ro

through the 2 mm thick polyethylene body of the sampler (Fauvelle et al., 2017b). Other samplers have also exploited the same concept but with a different approach. For example, the o-DGT

-p

sampler achieved this by using a 0.8 mm thick diffusive gel layer (Chen et al., 2012) while Martin et

re

al. (2003) used a ceramic tube.

The overall aim of this study then was to develop, calibrate and validate a MPT passive sampler

lP

for quantitative estimation of licit, illicit drugs and PPCPs by in-situ deployment in wastewater

na

influent for use with WBE. To achieve this, the following specific objectives were: (i) development and calibration of two configurations (Phenomenex Strata-X solid phase extraction

ur

(SPE) sorbent and Strata-X suspended in agarose gel sorbent) of the MPT passive sampler; (ii) determination and modelling of chemical-specific uptake and sampling rates of drugs and PPCPs

Jo

of interest in the MPT and

(iii) validation of the use of MPTs in WWTP influent by trialling them in an independent deployment where water concentration estimates were compared from simultaneous MPT and high-resolution time proportional composite samples.

2. Materials and Methods 2.1 Solvents and Chemicals All equipment and materials (including glassware) were thoroughly rinsed with acetone (ACE) and methanol (MeOH) (Merck, Darmstadt, Germany; purity 99.8%) and allowed to dry prior to use. Water with resistivity > 18.2 MΩ.cm (MilliQ) was obtained from a Millipore system. Samples were

4

Journal Pre-proof analysed for 30 illicit drugs and PPCPs, along with their metabolites (Table S1). All analytical standards were purchased from Novachem (VIC, Australia) and the 21 mass-labelled internal and surrogate standards are also listed in Table S1. A mixed working stock solution (1000 ng mL-1) of the surrogate standards was prepared in MeOH. Bulk Strata-X sorbent (100 g) was purchased from Phenomenex (Lane Cove, NSW, Australia). Additionally, agarose gel was molecular biology grade/DNA grade, purchased from Progen BioSciences (Brisbane, QLD, Australia).

2.2 Sample Preparation The assembly of the MPT samplers has been described in Fauvelle et al. (2017b). The MPTs

of

themselves (12 mm O.D., 35% porosity, 2.5 μm pore size) were purchased from Pall Corp. (Crailsheim, Germany). The tubes were cut into 4 cm long cylinders (10 cm2 inner surface area) and

ro

cleaned in MeOH for 24 hours on a shaker, followed by ultra-pure water (MilliQ) for 24 hours. They

(i)

-p

were then left to dry in a closed fumehood. Two MPT configurations were investigated in this study: MPT containing Strata-X. A plastic tubing cap (5 mm; Stockcap, Australia) was placed on

re

one end of the 4 cm MPT tube, then 400 ± 10 mg of Strata-X SPE sorbent (Phenomenex,

lP

Melbourne, Australia) was weighed into the tube, which was then capped on the other end. Filled tubes were conditioned in MeOH on a shaker for 24 hours followed by MilliQ water for another 24 hours. The MilliQ water was replaced and samplers were kept

(ii)

na

submerged and refrigerated (4°C) until deployment. MPT containing agarose and Strata-X. 10 g of Strata-X sorbent was conditioned and

ur

hydrated, firstly with 200 mL of MeOH and then with 200 mL of MilliQ water, using a

Jo

Buchner funnel and vacuum flask. The conditioned Strata-X sorbent was sonicated with MilliQ water and then combined with 200 mg of agarose powder in a single-use 15 mL polyethylene Falcon tube. This tube was manually shaken and immediately placed into a hot water bath at 90°C for 10 minutes, while shaking occasionally. The mixture was cast into borosilicate glass tubes (8 mm I.D., 12 mm O.D.), capped using plastic tubing caps (mentioned above) and submerged into an ice bath for 5 minutes. Tubes were then placed in the refrigerator (4°C) for 2 hours until the gel set. Following this, the glass tubes were placed on a clean bench covered in aluminium foil. Gels were allowed to slide out of the glass tubes onto the clean aluminium foil and were cut with a stainlesssteel knife to length (one glass tube yields gels for approximately 5 MPTs). Gels were finally inserted into the MPT tubes whilst submerged in a MilliQ water water bath. This was done to eliminate air bubbles in the MPT tubing during assembly (Fauvelle et al.,

5

Journal Pre-proof 2017b). Samplers were capped and kept submerged in MilliQ water in the refrigerator (4°C) until removed for deployment.

2.3 WWTP Study Design MPTs were deployed for 4, 7, 15, 21 and 29 days in the influent stream of a municipal Wastewater Treatment plant (WWTP) located in South East Queensland, Australia. The site treats primarily municipal sewage for a population of approximately 200,000 people (Australian Census 2016). Both MPT configurations containing Strata-X alone and Strata-X + agarose gel were deployed

of

in triplicate, mounted on pre-cleaned plastic cages and fixed in place with pre-cleaned zip ties (Figure S1). Samplers were deployed from the 12th of January 2017 (day 1) to the 10th of February

ro

(day 29) and were removed in a staggered pattern. Composite 24-hour grab samples (500 mL) were collected on days 6, 12, 20 and 26 during the passive deployment using a refrigerated autosampler.

-p

All samples were transported to the laboratory immediately after collection and stored in a freezer (-

re

20 °C) until the time of extraction. For the validation phase of the work, triplicate MPT samplers containing Strata X + agarose gel were deployed for 7 days at the same WWTP during the week of

lP

the 2016 Australian Census (August 2016). This was an opportunity to link levels of licit, illicit drugs and PPCPs to WWTP catchment area population data for WBE purposes. Samples were kept at -20

ur

2.4 Sample Extraction

na

°C until extraction and analysis as described below.

Jo

2.4.1 MPT passive sampler extraction procedure Individual MPT samplers from calibration and validation phases of the work were removed from their deployment cages, and any debris left on the samplers removed using sterile laboratory wipes. Each sampler was then placed into separate labelled, single-use 15 mL polypropylene Falcon tubes. Blank MPT tubes (n = 6) and solvent blanks (n = 1) were extracted alongside samples. Samples in the tube were spiked with 20 μL of a 500 ng mL-1 isotopically labelled illicit drug surrogate standard and 20 μL of a 1000 ng mL-1 isotopically labelled PPCP surrogate standard, by removing the cap of the MPT on one side, spiking the solutions and then closing the cap. MeOH (4 mL) was added and the Falcon tube sonicated for 10 minutes. This was repeated 4 times, transferring and combining the solutions into a clean Falcon tube after each extraction step. Extracts were then evaporated to 1mL under a gentle stream of nitrogen and centrifuged for 10 minutes at 2500 rpm (22°C). The supernatant was then transferred into a 2 mL vial, where it was further reduced under nitrogen to 200 μL. Samples were then made up to a final volume of 1 mL with MilliQ water. 6

Journal Pre-proof 2.4.2 Composite sample extraction procedure Solid phase extraction was used to extract the composite grab samples, based on the method of van Nuijs (2009) with minor changes, as described below. Composite samples were defrosted and a 100 mL aliquot of the sample centrifuged. The supernatant was removed (> 99.5 mL) and placed into a clean bottle. Then 5 ng of isotope-labelled illicit drug and PPCP surrogate standards were spiked into each sample prior to extraction. One blank (i.e. MilliQ water) sample was added with each batch of composite samples being extracted. SPE cartridges (6 cm3; 200mg) (Strata-X, Phenomenex) were conditioned first with 4 mL of MeOH and then 4 mL of MilliQ water. All samples were then loaded onto SPE cartridges and eluted with 2 x 4 mL of MeOH on a SPE vacuum manifold. Finally, samples

of

were concentrated under a gentle nitrogen stream to 200 μL and made up to 1mL with MilliQ water.

ro

2.5 Analysis

-p

Samples were analysed using Liquid Chromatography (LC) (Shimadzu, Nexera HPLC system, Kyoto, Japan) coupled with tandem electrospray ionization (ESI) mass spectrometry (MS/MS) (SCIEX

re

API 5500 and 6500+ QTRAP® Mass Spectrometers, Ontario, Canada). Both positive and negative ionization modes were operated based on a scheduled multiple reaction monitoring (MRM)

lP

switching process. Table S2 summarizes the chromatographic conditions used. The isotopic ratio between internal standard and analyte was used to quantify target analytes. Analytes that did not

na

have matching mass-labelled standards used external calibrations or surrogate standards. A sevenpoint calibration of native standards (within the range of 0.1 – 40 ng mL-1) was prepared for

Jo

(Sciex).

ur

quantitation of analytes. Data processing was undertaken using MultiQuant and Analyst Software

2.6 Quality Assurance and Quality Control (QA/QC) Laboratory passive sampler blanks (n = 6) as well as laboratory composite blanks (n = 1) were prepared, extracted and analysed alongside the passive samplers. In the MPTs, blank levels in samples were less than 1% of the mass accumulated in passive samplers at day 4 (t = 1) (Table S1). Triplicate MPT sampler deployment also enabled evaluation of reproducibility within deployment. Acceptable reproducibility (expressed as CV%) was observed (≤ 26% for MPTs with Strata-X and ≤ 33% for MPTs with Strata-X + agarose gel (Table 1). The only PPCPs or illicit drugs that were detected in the composite blanks were acesulfame and carbamazepine (2.50 and 0.11 ng mL-1, respectively) (Table S1). Limits of detection were determined by calculating three times the standard deviation (SD) from either a low calibration standard or from

7

Journal Pre-proof average blank levels, if blank contamination occurred. Instrument quantitation (LOQ) limits were taken as 3 times the LOD. Recovery of chemicals was verified by spiking labelled surrogate internal standards into all passive samplers and composites as well as QA/QC samples prior to the extraction process. Nonextracted side spikes were prepared alongside calibration samples and represent 100% recovery when used to calculate recovery of calibration samples. Recoveries for PPCPs were ≥ 20% for the majority of analytes, with notably higher recoveries for the agarose sampler (with ≥ 50% for the majority of analytes). Methods for quantification of drugs have been well established in our laboratory and we routinely participate in the SCORE and passive sampling interlaboratory studies

of

(Hageman et al., 2019; Vrana et al., 2016). Laboratory performance is also routinely assessed

ro

through analysis of intra-laboratory QA/QC samples (aliquoted pooled wastewater samples included

-p

in each batch).

Serial dilutions were also conducted to confirm whether analyses of analytes were influenced by

re

matrix suppression, which is a known issue with ESI and LC-MS/MS (e.g. Gonzalez-Marino et al., 2012). Matrix effects were not apparent for the majority of quantified analytes with MPTs

lP

containing Strata-X. Those analytes that did exhibit effects were drugs and PPCPs that were neutral or largely so, rather than ionized in solution (Figures S2 and S3). For approximately half of the

na

analytes with Strata-X/agarose gel samplers however, results indicated matrix effects were encountered during analysis but the occurrence of matrix effects and signal suppression was

Jo

2.7 Data Modelling

ur

analyte/sorbent specific and did not display any discernible patterns.

The accumulation of polar organic chemicals in the sorbent phase of a passive sampling device can be described using a one compartment, 1st order kinetic model (Equation 1) where NS is the amount of the compound in the sampler sorbent (ng sampler-1), KSW the sorbent-water partition coefficient (mL g-1), CW the concentration of the compound in water (ng mL-1), RS the sampling rate (mL d-1), t the time deployed (days) and mS the mass of the sorbent (g).

𝑁𝑆 /𝐶𝑊 = 𝐾𝑆𝑊 . 𝑚𝑆 [1 − 𝑒

8

𝑅𝑆. 𝑡 ) 𝑚𝑠 .𝐾𝑆𝑊

−(

]

(1)

Journal Pre-proof When uptake of chemicals into the sampler is linear with time i.e. the sampler is operating in kinetic mode, accumulation can be reduced to a linear approximation model (Equation 2) (Huckins et al., 2006).

𝑁𝑆 ⁄𝐶 = 𝑅𝑆 . 𝑡 𝑊

(2)

Values of Rs for linear uptake were estimated from Equation 2 by unweighted linear regression and for analytes in the curvilinear stage of accumulation, from Equation 1 by unweighted nonlinear least

of

squares regression using GraphPad Prism V7.03 (San Diego, CA, USA, 2017).

ro

3. Results and discussion

-p

3.1 Uptake of target compounds

Thirty target compounds from WWTP influent were detected in MPT samplers of both

re

configurations and 29 were detected and quantified in extracted parallel composite grab samples (Table 1). Detected analytes include the PPCPs atenolol, atorvastatin, carbamazepine, citalopram,

lP

codeine, cotinine, furosemide, hydrochlorothiazide, hydroxycotinine, iopromide, naproxen, nicotine, oxycodone, paracetamol, temazepam, tramadol and venlafaxine. The following illicit drugs were also

na

detected: amphetamine, benzoylecgonine, EDDP, MDMA, methamphetamine and morphine. The analyte mass accumulated in the MPT (Ns) ranged over approximately five orders of magnitude from

Jo

respectively.

ur

0.31 ng sampler-1 to 9.8 µg sampler-1 for desmethyldiazepam (Strata-X) and paracetamol (Strata-X),

Cocaine was detected only in the MPT samplers but not in the composite grab samples (LOQ 0.01 ng mL-1) and thus sampling rates could not be determined. The linear accumulation demonstrated in the passive samplers during calibration (R2 = 0.93 and 0.87 (Strata-X + agarose and Strata-X respectively)) suggests however that these samplers can also be successfully deployed for trace levels of cocaine following calibration. Target analytes were categorized into three types based on their temporal MPT uptake behaviour i.e. (i) chemicals demonstrating linear uptake over the 29 day deployment period, (ii) chemicals demonstrating a curvilinear uptake and (iii) chemicals attaining equilibrium or alternatively exhibiting no uptake or no discernible uptake pattern. The uptake of PPCPs and drugs was assessed as linear when linear regression analysis of the ratio of analyte mass accumulated to

9

Journal Pre-proof aqueous concentration (Ns/Cw) plotted against deployment time showed R2 values greater than 0.84 (Kaserzon et al., 2019). The substances that fulfilled this criterion are shown in Table 1.

3.2 Linear Uptake Twenty of the analytes quantified in extracted composite grab samples showed linear uptake over the 29 day calibration MPT deployment period with at least one of the sampler configurations, exhibiting R2 values ranging from 0.84 – 0.99 (Table 1). Derived sampling rates ranged from 0.25 to 16.9 mL d-1 for the Strata-X equipped sampler (for ibuprofen and verapamil, respectively) and 0.58 to 9.1 mL d-1 for the Strata-X + agarose sampler (for ibuprofen and methamphetamine, respectively).

of

Of these 20 analytes demonstrating linear uptake, 11 had no significant difference between slopes for the two MPT sampler configurations (slope ratio (expressed as %) of > 70% (Strata X / Strata X +

ro

agarose)). This group comprised atenolol, amphetamine, benzoylecgonine, carbamazepine, codeine, hydrochlorothiazide, MDMA, methamphetamine, morphine, nicotine and temazepam (Figure 1). The

-p

linear uptake with time of target analytes into the MPT indicates the samplers were operating in

re

kinetic mode. Consequently, Rs was deduced based on the entire deployment period for calibration and the data used for subsequent validation. Furthermore, similar sampling rates with Strata-X and

lP

Strata-X + agarose implies that these sorbent phases are not controlling uptake, rather the MPT itself is largely responsible for this with these compounds (Kaserzon et al., 2019).

na

Related studies employing POCIS containing either “pharmaceutical” (i.e. a hydroxylated polystyrene–divinylbenzene polymer and activated carbon dispersed on styrene divinylbenzene

ur

copolymer beads) or “pesticide” (i.e. Oasis HLB) sorbents or o-DGT samplers with XAD-18 resin (Guo et al., 2017) in WWTPs have resulted in derived sampling rates for illicit drugs and PPCPs generally 2

Jo

to 7 times higher than those in this work when normalised to surface area (Table 2). It has been widely acknowledged that analyte uptake in the POCIS sampler is often controlled by the WBL, consequently sampling rates are often a product of unpredictable and fluctuating water flow rate such as might occur in a WWTP environment (Alvarez et al., 2004; Fauvelle et al., 2017b; Harman et al., 2011; Li et al., 2011b). Alvarez et al. (2004) reported increases in sampling rate of up to an order of magnitude could occur with changes in water velocity. With DGT samplers, the agarose gel of the diffusive layer controls the sampling rate, which reduces the influence of varying hydrodynamic conditions (Belles et al., 2017; Guo et al., 2017). Fauvelle et al. (2017b) suggest that the MPT minimises these effects by uptake being largely governed by diffusion through the microporous polyethylene. If the MPT sampler is not under WBL control, then in-situ calibration under various flow rate and/or turbulence conditions may not be required. Consequently, observed 10

Journal Pre-proof differences between sampling rates shown in Table 2 are likely the result of different controlling factors on analyte diffusion and uptake with all the various sampler (configurations). The kinetic mode of sampling, characterized by linear uptake with time, can be taken as the difference in time between the start of exposure to the half-time to equilibrium (Townsend et al., 2018). Higher sampling rates therefore decrease the duration of this period and the time to equilibrium. For example, several analytes including the cocaine metabolite benzoylecgonine have been reported to attain equilibrium after 14 days with POCIS deployed in a WWTP (Harman et al., 2011). In contrast, the MPT sampler demonstrated decreased sampling rates compared to other samplers (Table 2). This means it takes longer to accumulate a given amount of analyte, an

of

important consideration with those at trace or ultra-trace levels, but half-times are extended. In this

ro

work, linear uptake was observed for a suite of significant illicit and licit drugs such as amphetamine, methamphetamine, MDMA, benzoylecgonine, nicotine, carbamazepine and atenolol, none of which

-p

reached equilibrium during deployment periods of up to 29 days.

re

Not all licit, illicit drugs and PPCPs were accumulated in a linear manner with time in the MPT samplers of both configurations however. For example, the X-ray phase contrast agent iopromide

lP

has the largest molar mass of the target compounds (791.1 g mol-1) and displayed consistent linear uptake in the MPT containing Strata-X + agarose sorbent with a Rs of 1.4 mL d-1, whilst the Strata-X

na

equipped sampler showed a curvilinear uptake profile (Rs of 2.0 mL d-1), approaching effective equilibrium after 15 days. In contrast, citalopram, a substituted dihydrobenzofuran antidepressant,

ur

demonstrated linear uptake in the Strata-X sampler (Rs = 3.5mL d-1), but no pattern was discernible for the agarose-containing MPT (Figure 2). Apart from a difference in size, iopromide is anionic

Jo

whilst citalopram is cationic under the pH conditions of WWTP influent in this area (7.0 to 8.0) (Cardenas et al., 2016). The differential behaviour with the two sorbent configurations suggests the involvement of these phases, at least to some extent, in uptake and accumulation of these compounds with analyte size and/or charge perhaps playing a role. Any influence of microoroganisms also needs to be considered in a WWTP influent environment. Fauvelle et al. (2017b) noted that the MPT pore size (2.5 μm) is large enough to allow microorganisms to pass through the polyethylene shell. Therefore, microbial transformation of specific compounds or microbial effects on the agarose gel itself may occur depending on the prevailing conditions and sorbent type, providing discrimination between MPT configurations. In addition, the pore size may allow not only dissolved analytes, but those associated with colloidal and fine particles to traverse as well.

11

Journal Pre-proof 3.3 Curvilinear uptake A range of analytes exhibited curvilinear uptake behaviour with both MPT sampler configurations. RS data (Table 1) and KSW values (Table S3) are derived from use of Equation 1. A comparison of these KSW values with literature data (Belles et al., 2017) is shown in Figure S4. This group of analytes includes compounds present in the influent as anionic species e.g. atorvastatin and naproxen, neutral or largely neutral species e.g. cotinine, hydroxycotinine and paracetamol and largely cationic species e.g. oxycodone and venlafaxine (Figure S5). Non-linear kinetics may be a response to fluctuating levels of PPCPs and illicit drugs in the influent (Hawker, 2010). A reduction of sampling rates due to biofouling, and the presence of varying dissolved and suspended solids levels

of

affecting diffusive uptake also need to be considered (Li et al., 2011a). Biofouling on the surface of the MPT samplers was observed during their calibration and increased with deployment time.

ro

However, the occurrence of linear uptake observed for 19 target analytes (Strata-X) suggests that

-p

biofouling was not having a time varying effect on mass transfer kinetics during the 29 day

re

deployment period.

3.4 Equilibrium sampling

lP

The artificial sweetener acesulfame displayed evidence of rapid equilibration in both Strata-X and Strata-X + agarose samplers (Figure S6). Equilibration with ambient WWTP influent (9.9 ± 0.98

na

ng mL-1) concentrations occurred prior to the first MPT collection (t = 4 days). Given the amount accumulated in the Strata-X sampler, a Strata-X/water distribution ratio of 101.87 mL g-1 may be

ur

calculated. Belles et al. (2017) measured equilibrium Strata-X/water partition coefficients for various organic micropollutants and found them to range between 1.6 and 0.2 log units above their log K OW

Jo

values. Acesulfame is relatively water soluble and at pH= 7.5, log DOW = -0.55, but given the aqueous medium is not pure water but influent wastewater, the increased dissolved solids content would tend to increase the partitioning of acesulfame into the sorbent. Thus, the observed results in this work are likely consistent with equilibrium partitioning. This equilibration implies a rapid sampling rate for one of the most polar of the target compounds and provides more information of the operational characteristics of the MPT samplers but should be confirmed with further data.

3.5 Validation of MPT samplers deployed in WWTP influent A national wastewater monitoring study conducted simultaneously with the Australian Census of Population and Housing (2016) provided an opportunity to validate the MPT sampler’s performance for primarily illicit drugs and some PPCPs of interest (O'Brien et al., 2018). MPTs containing Strata-X + agarose gel (n = 3) were deployed for 7 days during this Census in influent of the same WWTP

12

Journal Pre-proof plant where the calibration was conducted. Sampling rates derived from that phase of the work were used to estimate concentrations from the amount of analyte accumulated in the samplers (Equation 2). MPT-derived data were then compared to high-resolution composite grab sampling data collected over the same 7 day period at the same location for those analytes that were quantified with both techniques in this deployment (Figure 3). Overall, a very good correlation (R2 = 0.95) was observed for a range of amphetamine-type stimulants, opioids and metabolites as well as nicotine. Concentrations spanned from approximately 0.1 ng mL-1 (benzoylecgonine) up to 6.0 ng mL-1 (nicotine). Most of these compounds were present as cations, with the anti-epileptic carbamazepine being neutral and the cocaine metabolite

ro

of

benzoylecgonine present as a zwitterion.

3.6 Sampler Characteristics and Limitations

-p

WWTPs are a challenging environment for passive samplers with varying flow rates and turbulence and relatively high levels of dissolved and suspended solids. The accumulation of illicit drugs and

re

PPCPs into the MPT sampler can be considered a multi-stage process with diffusion through a WBL

lP

and PE matrix or aqueous pores in the PE to the sorbent. The reciprocal of the overall mass transfer coefficient contains terms representing resistance to mass transfer in all these phases (Kaserzon et al., 2019). The strategy underpinning development of the MPT for this deployment application was

na

to reduce or eliminate the influence of the WBL whose thickness will fluctuate in a WWTP environment. If resistance in the PE matrix (2 mm thick) or aqueous pores is sufficiently great, this

ur

will control uptake. Reduced sampling rates compared to other passive sampler configurations e.g. POCIS, similar sampling rates for MPTs containing different sorbents and successful validation using

Jo

independently-derived sampling rates all indicate the success of this strategy. The target analytes included compounds that are neutral, zwitterionic, cationic and anionic under prevailing conditions. Li et al. (2011a) found that diffusion across the polyethersulfone membrane (0.1 μm pore size) of POCIS was greatest for neutral forms and found positive linear relationships between RS and the pH-dependent octanol/water distribution ratio (DOW). The preliminary data from this current work also suggests sampling rates are greater for neutral species than ionized ones. When plotted against log DOW (pH = 7.4 and 298 K; (ACD/Labs) Software V11.02 (© 1994-2019 ACD/Labs)), the data from the MPT samplers with Strata-X in particular tend to bifurcate into two groups based on charge, which appears to be a primary discriminant. However, linear regression analysis shows slopes are not significantly different from zero (Figure S7 and Table S4). While caution should be exercised given the relatively small number of compounds involved, this suggests 13

Journal Pre-proof the hydrophobicity of these drugs and PPCPs is not a principal factor in determining the magnitude of RS. Further, non-target analytes may be able to be quantified with reasonable accuracy simply based on charge within the log DOW range of calibrants (-1.85 to 4.44). However, if more hydrophobic compounds are targeted, there is a possibility of more extensive partitioning to the MPT tube rather than the Strata-X sorbent (Kaserzon et al., 2014; Vermeirssen et al., 2012). While mass transfer through the MPT matrix or aqueous pores may be the limiting process for uptake, the sorbent is the ultimate repository. Strata-X is one of a number of commercially available functionalized polymeric materials with affinity for acidic, neutral and basic sorbates (Weigel et al., 2004). This is due to multiple modes of retention; necessary when working with illicit drugs and

of

PPCPs of different structure and acid-base speciation in wastewater influent. Potential retention

ro

modes include 𝜋 − 𝜋 bonding, hydrogen bonding and other dipole-dipole attractions as well as hydrophobic interactions. There is no apparent facility for coulombic interactions as would occur

-p

with ion-exchange sorbents. Approximately two-thirds of the target compounds found in the wastewater could be linearly accumulated in the MPT samplers over a deployment period of up to

re

29 days. There were relatively few negatively charged conjugate bases of acidic substances present

lP

though, so a fuller demonstration and understanding of the capability of these devices is required. In addition, the effects of biofouling, changing temperature and chemical stability (such as with

na

cocaine) may compromise quantification efforts and need to be further investigated. There is also scope for sampler performance optimization in regard to matrix effects on

ur

analysis. LC-MS/MS is often the analytical technique employed for quantification of polar or ionized contaminants such as drugs and PPCPs (e.g. Baz-Lomba et al., 2017), and ESI interfaces often provide

Jo

good sensitivity. However with complex matrices such as wastewater, the presence of endogenous material such as dissolved organic matter can supress ionization, meaning signal response from matrix samples and standards can vary (Babic et al., 2010). Matrix effects were not evident for all target compounds, but did seem to affect some neutral drugs and PPCPs such as carbamazepine (Figures S2 and S3). For those target substances that were not accumulated satisfactorily with the MPT sampler, their detection may still be useful for qualitative purposes and future sampler modifications and calibrations may provide better resolution for these challenging analytes.

4. Conclusions In this work, the applicability of MPT passive samplers with a multi-mode sorbent as a tool for detection and quantitative estimation of priority drugs of abuse and PPCPs in WWTP influent was 14

Journal Pre-proof examined and validated. Drugs of high use and concern in communities such as methamphetamine, MDMA, nicotine and morphine showed linear uptake over a calibration period of 29 days. Sampling rates of drugs and PPCPs of interest demonstrating linear accumulation with time spanned from 0.25 mL d-1 to 16.9 mL d-1 for ibuprofen and verapamil, respectively and MPT LODs were within the range of 0.003 to 0.84 ng mL-1 for drugs such as cocaine and ibuprofen, respectively. This suggests sufficient sensitivity for deployments in WWTP influent systems. The MPT sampler was able to effectively determine concentrations of PPCPs and illicit drugs in wastewater influent (R2 = 0.95) in a separate independent validation deployment using sampling rates derived in the calibration phase of this work. While further investigation into the MPT performance in other environmental matrices,

of

such as surface water, groundwater and WWTP effluent would be beneficial to establish its wider applicability, the sampler shows promise as a device for providing essential monitoring data for

ro

WBE.

-p

Acknowledgements

re

The authors would like to acknowledge and thank Tim Reeks, Geoff Eaglesham, Sean van Niekerk, Rory Verhagen and the team at QAEHS for their support and guidance during this study.

lP

The authors would also like to thank the staff at the WWTP for their assistance and contributions

na

during the calibration and validation studies.

Jo

References

ur

Supplementary data is available

Alvarez DA, Petty JD, Huckins JN, Jones-Lepp TL, Getting DT, Goddard JP, et al. Development of a passive, in situ, integrative sampler for hydrophilic organic contaminants in aquatic environments. Environmental Toxicology and Chemistry 2004; 23: 1640-1648. Babic S, Mutavdzic Pavlovic D, Asperger D, Perisa M, Zrncic M, Horvat AJ, et al. Determination of multi-class pharmaceuticals in wastewater by liquid chromatography-tandem mass spectrometry (LC-MS-MS). Anal Bioanal Chem 2010; 398: 1185-94. Baz-Lomba JA, Harman C, Reid M, Thomas KV. Passive sampling of wastewater as a tool for the longterm monitoring of community exposure: Illicit and prescription drug trends as a proof of concept. Water Res 2017; 121: 221-230. Belles A, Alary C, Aminot Y, Readman JW, Franke C. Calibration and response of an agarose gel based passive sampler to record short pulses of aquatic organic pollutants. Talanta 2017; 165: 1-9. Booij K, Vrana B, Huckins JN. Chapter 7 Theory, modelling and calibration of passive samplers used in water monitoring. In: Greenwood R, Mills G, Vrana B, editors. Comprehensive Analytical Chemistry. 48. Elsevier, 2007, pp. 141-169.

15

Journal Pre-proof

Jo

ur

na

lP

re

-p

ro

of

Cardenas MA, Ali I, Lai FY, Dawes L, Thier R, Rajapakse J. Removal of micropollutants through a biological wastewater treatment plant in a subtropical climate, Queensland-Australia. J Environ Health Sci Eng 2016; 14: 14. Castiglioni S, Bijlsma L, Covaci A, Emke E, Hernandez F, Reid M, et al. Evaluation of Uncertainties Associated with the Determination of Community Drug Use through the Measurement of Sewage Drug Biomarkers. Environmental Science & Technology 2013; 47: 1452-1460. Chen CE, Zhang H, Jones KC. A novel passive water sampler for in situ sampling of antibiotics. J Environ Monit 2012; 14: 1523-30. Choi PM, Tscharke BJ, Donner E, O'Brien JW, Grant SC, Kaserzon SL, et al. Wastewater-based epidemiology biomarkers: Past, present and future. TrAC Trends in Analytical Chemistry 2018; 105: 14. Di Carro M, Bono L, Magi, E. A simple recirculating flow system for the calibration of polar organic chemical integrative samplers (POCIS): effect of flow rate on different water pollutants. Talanta 2014; 120: 30-33. Fauvelle V, Kaserzon SL, Montero N, Lissalde S, Allan IJ, Mills G, et al. Dealing with Flow Effects on the Uptake of Polar Compounds by Passive Samplers. Environ Sci Technol 2017a; 51: 25362537. Fauvelle V, Montero N, Mueller JF, Banks A, Mazzella N, Kaserzon SL. Glyphosate and AMPA passive sampling in freshwater using a microporous polyethylene diffusion sampler. Chemosphere 2017b; 188: 241-248. Gonzalez-Marino I, Quintana JB, Rodriguez I, Gonzalez-Diez M, Cela R. Screening and selective quantification of illicit drugs in wastewater by mixed-mode solid-phase extraction and quadrupole-time-of-flight liquid chromatography-mass spectrometry. Anal Chem 2012; 84: 1708-17. Gonzalez-Rey M, Tapie N, Le Menach K, Devier MH, Budzinski H, Bebianno MJ. Occurrence of pharmaceutical compounds and pesticides in aquatic systems. Mar Pollut Bull 2015; 96: 384400. Guo C, Zhang T, Hou S, Lv J, Zhang Y, Wu F, et al. Investigation and Application of a New Passive Sampling Technique for in Situ Monitoring of Illicit Drugs in Waste Waters and Rivers. Environ Sci Technol 2017; 51: 9101-9108. Hageman KJ, Aebig CHF, Luong KH, Kaserzon SL, Wong CS, Reeks T, et al. Current-Use Pesticides in New Zealand Streams: Comparing Results from Grab Samples and Three Types of Passive Samplers. Environmental Pollution 2019; 254(Pt A): 112973. Harman C, Allan IJ, Vermeirssen EL. Calibration and use of the polar organic chemical integrative sampler--a critical review. Environ Toxicol Chem 2012; 31: 2724-38. Harman C, Reid M, Thomas KV. In Situ Calibration of a Passive Sampling Device for Selected Illicit Drugs and Their Metabolites in Wastewater, And Subsequent Year-Long Assessment of Community Drug Usage. Environmental Science & Technology 2011; 45: 5676-5682. Hawker DW. Modeling the response of passive samplers to varying ambient fluid concentrations of organic contaminants. Environ Toxicol Chem 2010; 29: 591-6. Huckins JN, Booij K, Petty JD. Monitors of Organic Chemicals in the Environment: Semipermeable Membrane Devices. Boston, MA: Springer US, 2006. Jones-Lepp TL, Alvarez DA, Petty JD, Huckins JN. Polar Organic Chemical Integrative Sampling and Liquid Chromatography–Electrospray/Ion-Trap Mass Spectrometry for Assessing Selected Prescription and Illicit Drugs in Treated Sewage Effluents. Archives of Environmental Contamination and Toxicology 2004; 47: 427-439. Kaserzon SL, Hawker DW, Kennedy K, Bartkow M, Carter S, Booij K, et al. Characterisation and comparison of the uptake of ionizable and polar pesticides, pharmaceuticals and personal care products by POCIS and Chemcatchers. Environ Sci Process Impacts 2014; 16: 2517-26. Kaserzon SL, Vijayasarathy S, Braunig J, Mueller L, Hawker DW, Thomas KV, et al. Calibration and validation of a novel passive sampling device for the time integrative monitoring of per- and 16

Journal Pre-proof

Jo

ur

na

lP

re

-p

ro

of

polyfluoroalkyl substances (PFASs) and precursors in contaminated groundwater. J Hazard Mater 2019; 366: 423-431. Kingston JK, Greenwood R, Mills GA, Morrison GM, Persson LB. Development of a novel passive sampling system for the time-averaged measurement of a range of organic pollutants in aquatic environments. Journal of Environmental Monitoring 2000; 2: 487-495. Li HX, Helm PA, Paterson G, Metcalfe CD. The effects of dissolved organic matter and pH on sampling rates for polar organic chemical integrative samplers (POCIS). Chemosphere 2011a; 83: 271280. Li HX, Vermeirssen ELM, Helm PA, Metcalfe CD. Controlled field evaluation of water flow rate effects on sampling polar organic compounds using polar organic chemical integrative samplers. Environmental Toxicology and Chemistry 2011b; 29: 2461-2469. Martin H, Patterson BM, Davis GB, Grathwohl P. Field trial of contaminant groundwater monitoring: comparing time-integrating ceramic dosimeters and conventional water sampling. Environmental science & technology, 2003; 37: 1360-1364. Metcalfe CD, Beddows PA, Bouchot GG, Metcalfe TL, Li HX, Van Lavieren H. Contaminants in the coastal karst aquifer system along the Caribbean coast of the Yucatan Peninsula, Mexico. Environmental Pollution 2011; 159: 991-997. Morin N, Miège C, Coquery M, Randon J. Chemical calibration, performance, validation and applications of the polar organic chemical integrative sampler (POCIS) in aquatic environments. TrAC Trends in Analytical Chemistry 2012; 36: 144-175. O'Brien JW, Grant S, Banks APW, Bruno R, Carter S, Choi PM, et al. A National Wastewater Monitoring Program for a better understanding of public health: A case study using the Australian Census. Environment International 2018; 122: 12. Salgado R, Marques R, Noronha JP, Mexia JT, Carvalho G, Oehmen A, et al. Assessing the diurnal variability of pharmaceutical and personal care products in a full-scale activated sludge plant. Environ Pollut 2011; 159: 2359-67. Townsend I, Jones L, Broom M, Gravell A, Schumacher M, Fones GR, et al. Calibration and application of the Chemcatcher(R) passive sampler for monitoring acidic herbicides in the River Exe, UK catchment. Environ Sci Pollut Res Int 2018; 25: 25130-25142. van Nuijs AL, Tarcomnicu I, Bervoets L, Blust R, Jorens PG, Neels H, et al. Analysis of drugs of abuse in wastewater by hydrophilic interaction liquid chromatography-tandem mass spectrometry. Anal Bioanal Chem 2009; 395: 819-28. Vermeirssen EL, Dietschweiler C, Escher BI, van der Voet J, Hollender J. Transfer kinetics of polar organic compounds over polyethersulfone membranes in the passive samplers POCIS and Chemcatcher. Environ Sci Technol 2012; 46: 6759-66. Vrana B, Mills GA, Dominiak E, Greenwood R. Calibration of the Chemcatcher passive sampler for the monitoring of priority organic pollutants in water. Environ Pollut 2006; 142: 333-43. Vrana B, Smedes F, Prokes R, Loos R, Mazzella N, Miege C, et al. An interlaboratory study on passive sampling of emerging water pollutants. Trends in Analytical Chemistry 2016; 76. Weigel S, Kallenborn R, Hühnerfuss H. Simultaneous solid-phase extraction of acidic, neutral and basic pharmaceuticals from aqueous samples at ambient (neutral) pH and their determination by gas chromatography–mass spectrometry. Journal of Chromatography A 2004; 1023: 183-195. Zuccato E, Chiabrando C, Castiglioni S, Calamari D, Bagnati R, Schiarea S, et al. Cocaine in surface waters: a new evidence-based tool to monitor community drug abuse. Environmental Health: A Global Access Science Source 2005; 4: 14.

17

Journal Pre-proof Declaration of interests

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

na

lP

re

-p

ro

of

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

Jo

ur

Dr Sarit Kaserzon, The University of Queensland (QAEHS)

18

Journal Pre-proof

Figure 1: Linear uptake of some of the illicit drugs and PPCPs by Strata-X + agarose (blue) and Strata-X (red) containing MPT passive samplers exhibiting similar sampling rates (slopes) over the 29-day calibration period (x axis). Mean WWTP influent concentrations from composite grab samples over this period are also shown (right y axis (triangles)).

Figure 2: Contrasting uptake behaviour of Iopromide and citalopram with the MPT Strata-X + agarose (blue) and Strata-X (red) MPT samplers over the 29-day deployment depicted. Also shown are concentrations from parallel composite grab sampling (right y axis (triangles)).

Figure 3: Composite grab sample water concentrations (n=2) (ng/mL) (y axis) for a range of illicit drugs and PPCPs (legend) plotted against MPT derived water concentrations (Cw; n=3) (ng/mL) (x axis), with compounds having smaller

of

concentrations depicted in the inset) (top right). The 1:1 relationship is represented by the dotted blue line and is

-p

ro

presented for comparison purposes.

re

Table 1: Wastewater treatment plant influent concentrations of illicit drugs and PPCPs from extracted composite -1

-1

grab samples (ng mL ; mean and standard deviation (SD)), mean derived Sampling Rates (Rs (mL d ) and their coefficient 2

lP

of variation (CV%) together with goodness of fit to linear and curvilinear uptake models (characterized by R ) of chemical uptake into MPT passive samplers containing the Strata-X and Strata-X + agarose sorbent phases. Sampling

na

rates (Rs) for linear analytes were derived from Equation 2, while Rs values for curvilinear analytes were derived from

ur

Equation 1. Also shown is acid/base speciation of these compounds at pH = 7.5.

Speciation @ pH 7.5

Acesulfame

Negative (-)/ Positive (+) Neutral (N) / Zwitterion (±)

Jo

Analyte

Water Concentration (ng mL-1)

Mean (n = 4)

SD

MPT passive sampler with Strata-X sorbent RS (mL d-1) (n = 3)

Linearity

MPT p

R2

CV% (ng/sampler) (n = 3)

RS (mL =

(-)

9.9

0.98

NA

Equilibrium

NA

4.0

N

Amphetamine

(+)

0.23

0.03

5.8 ± 0.37

Linear

0.98

7.6

7.0 ±

Atenolol

(+)

1.6

0.21

3.0 ± 0.38

Linear

0.94

15

3.2 ±

Benzoylecgonine

(±)

0.05

0.009

5.1 ± 0.43

Linear

0.97

11

5.0 ±

Carbamazepine

N

0.69

0.13

3.6 ± 0.25

Linear

0.98

11

4.1 ±

Citalopram

(+)

0.17

0.04

3.5 ± 0.30

Linear

0.97

21

N

Cocaine

(+)

N.D.

N.D.

Linear

0.87*

17

Codeine

(+)

2.3

0.19

2.9 ± 0.40

Linear

0.93

11

Desmethylcitalopram

(+)

0.04

0.008

7.4 ± 0.76

Linear

0.97**

26

EDDP

(+)

0.1

0.01

5.9 ± 0.89

Linear

0.92

17

8.7

Furosemide

(-)

1.5

0.35

1.7 ± 0.21

Linear

0.94

22

3.0 ±

Hydrochlorothiazide

N

1.3

0.25

2.3 ± 0.31

Linear

0.93

13

2.3 ±

Ibuprofen

(-)

14

2.3

0.25 ± 0.05

Linear

0.89

NA

0.58

19

3.2 ±

N

Journal Pre-proof (-)

2.2

0.31

2.0 ± 0.23

Curvilinear

0.97

24

1.4 ±

MDMA

(+)

0.03

0.008

7.6 ± 0.57

Linear

0.98

14

8.3 ±

Methamphetamine

(+)

1.3

0.08

8.7 ± 0.74

Linear

0.97

14

9.1

Morphine

(+)

1

0.1

6.8 ± 0.41

Linear

0.99

14

7.5 ±

Nicotine

(+)

3.6

0.36

5.3 ± 0.53

Linear

0.96

13

4.6 ±

Temazepam

N

0.3

0.024

2.9 ± 0.41

Linear

0.94

18

3.7 ±

Tramadol

(+)

0.45

0.07

2.0 ± 0.44

Linear

0.84

41

7.2 ±

Verapamil

(+)

0.02

0.003

16.9 ± 2.06

Linear

0.96**

24

N

Atorvastatin

(-)

0.2

0.02

9.6 ± 0.05

Curvilinear

0.94

24

N

Cotinine

N

4.2

0.34

6.5 ± 0.08

Curvilinear

1.00

11

6.5 ±

Desmethyldiazepam

N

0.013

0.001

8.9 ± 0.03

Curvilinear

0.99

13

N

Fluoxetine

(+)

0.03

0.003

NA***

Curvilinear

NA***

25

N

Hydroxycotinine

N

6.2

0.55

3.4 ± 0.10

Curvilinear

0.97

16

N

Naproxen

(-)

2.7

0.26

N

Oxycodone

(+)

0.1

0.004

16.6 ±

Paracetamol

N

34

Venlafaxine

(+)

1.1

of

Iopromide

0.97

14

11.7 ± 0.05

Curvilinear

1.00

16

42.0 ± 0.03

Curvilinear

0.82

26

N

6.6 ± 0.10 * R2 for cocaine represents the slope of mass per sampler plotted against time as no cocaine was detected in composite grab samples, hence sampling rate could not be derived

Curvilinear

0.81

12

13.4 ±

ro

1.6 ± 0.18

Curvilinear

9.6

-p

0.14

** Excluding one outlier on day 6

N.D. not detected

Jo

ur

na

lP

NA not applicable (regression parameters were not satisfied)

re

*** could not be estimated due to mass in samplers significantly exceeding calibration range

20

Journal Pre-proof Table 2: Comparison of sampling rates reported using POCIS or o-DGT and those determined with MPTs in this -1

-2

study. All Rs data are normalised to surface areas of the respective samplers (mL d cm ).

Surface Area

Methamphetamine

Baz-Lomba et

Bartelt-Hunt

Guo et

Harman et al.

al. (2014)

al. (2017)

et al. (2009)

al. (2017)

(2011)

POCIS

POCIS

POCIS

DGT

POCIS

2

41 cm

2

3.1cm

Assumed 41 2

48 cm

2.7

0.54

cm

Amphetamine Benzoylecgonine

1.2

Citalopram

9.0

3.2 0.58

2

10 cm

2

2.1

0.87

0.88

2.0

0.58

0.70

1.8

0.50

0.50

2.5/2.0

0.76

0.83

0.36

0.41

4.9

lP na ur Jo 21

MPT

7.1

re

Atenolol

Agarose

7.9

-p

1.5

Strata-X

6.3

MDMA 6.3

48 cm

This Study

5.4

0.69

Carbamazepine

2

of

Sampler type

Fedorova et

ro

Compound

0.35 0.29

0.32

Journal Pre-proof Graphical abstract

Highlights Microporous polyethylene passive samplers deployed in WWTP influent for ≤ 29 days



Uptake of illicit drugs and PPCPs allowed derivation of sampling rates



Drug/PPCP levels in separate deployment predicted well with derived sampling rates



Sampler shows promise for providing necessary drug monitoring data in wastewater

Jo

ur

na

lP

re

-p

ro

of



22

Figure 1

Figure 2

Figure 3