Project no.
003956
Project acronym
NOMIRACLE
Project title
Novel Methods for Integrated Risk Assessment of
Cumulative Stressors in Europe
Instrument
IP
Thematic Priority
1.1.6.3, ‘Global Change and Ecosystems’
Topic VII.1.1.a, ‘Development of risk assessment
methodologies’
Deliverable reference number and title:
D.2.2.16 Report on the evaluation of new extraction materials and analytical
techniques for polar organic compounds such as pesticides and
pharmaceutical residues in soils and waste water
Due date of deliverable: 1 November 2009
Start date of project: 1 November 2004
Actual submission date: 14 September 2009
Duration: 5 years
Organisation name of lead contractor for this deliverable: CSICRevision [draft, 1, 2, …]:
Project co-funded by the European Commission within the Sixth Framework Programme (2002-2006)
Dissemination Level
PU
PP
RE
CO
Public
Restricted to other programme participants (including the Commission Services)
Restricted to a group specified by the consortium (including the Commission Services)
Confidential, only for members of the consortium (including the Commission Services)
X
1
Authors and their organisation:
María J. García Galán, CSIC
M. Silvia Díaz-Cruz, CSIC
Damià Barceló, CSIC
Thomas Alsberg, ITM
J. Magner, ITM
Deliverable no:
D.2.2.16
Nature: R
Status: Draft 1
Dissemination
level: PU
Date of delivery:
Date of publishing:
Reviewed by (name and period):
2
TABLE OF CONTENTS
Summary
1. General remarks
2. Extraction and purification of environmental samples.
2.1. Sulfonamides in natural water samples
2.1.1 Offline SPE
2.1.2 Online SPE
2.2. Triazines and triazine metabolites in soils and natural waters
2.3. Polar organic compounds in waste waters
3. RESULTS
3.1. Sulfonamides
3.1.1. Offline SPE-LC-QqLIT-MS
3.1.2. Online SPE-LC-QqLIT-MS
3.2. Triazines
3.3. Polar organic contaminants
Conclusions
References
Annex A
Annex B
Annex C
3
D.2.2.16 Report on the evaluation of a new extraction material and different
analytical techniques for polar compounds such as pesticides and
pharmaceutical residues in soils and environmental waters.
SUMMARY
This report gathers various parallel studies carried out in Sweden and Spain for the determination of
polar contaminants such as pharmaceutical residues and pesticides in different water matrices and soil
(only for pesticides).
In first place, two new analytical methods based on liquid chromatography–quadrupole linear ion trap–
mass spectrometry (LC–QqLIT–MS) were developed for the determination of sulfonamide antibiotics and
one N4-acetylated metabolite in environmental waters. The first of them was based on offline solid phase
extraction, and special emphasis was devoted to the elimination of matrix components, evaluating three
different extraction/purification strategies: single cartridges (Oasis HLB and Oasis MCX) and tandem
(TD) extraction (combination of both). The second was based on on-line solid-phase extraction, and HLB
cartridges were employed.
In a parallel study, a new sensitive and selective method for the environmental trace analysis of triazines
and triazine metabolites was developed. For this purpose, commercially available molecularly imprinted
polymers (MIPS) were used fo the extraction and preconcentration of water and soils, followed by liquid
chromatography-tandem mass spectrometric (LC-MS/MS) analysis. The suitability of the method was
demonstrated through the analysis of several ground water and sludge-amended soils samples.
In the last work developed, a novel plastic material Poly(ethylene-co-vinyl acetate-co-carbon monoxide)
(PEVAC) was used as an absorptive passive equilibrium sampler and evaluated with regard to its ability
to assess the bioavailable concentration of polar organic compounds (POCs) in laboratory experiments
and in wastewater from a sewage treatment plant and compared with existing Solid-phase extraction
techniques (SPE).
1. General Remarks
Trace and ultra-trace analysis of polar organic contaminants in environmental samples generally requires
a pretreatment step to isolate and enrich the target analytes, and also to reduce the matrix interference
prior to chromatographic separation. Solid phase extraction (SPE) is usually the technique of choice for
4
the simultaneous extraction and concentration of many organic compounds present in aqueous samples
(1, 2, 3). Despite their attractive features, the classical SPE sorbents retain analytes by physicochemical
interactions that lead to co-extraction o f other matrix components and interfering substances, such as
fulvic and humic acids (4,5). Furthermore, when using liquid chromatography-mass spectrometry (LCMS) for analysis, signal suppression or enhancement occurs when matrix components are coeluted with
the analyte(s), especially using electrospray ionization (ESI) mode.
Molecularly imprinted polymers (MIPs) have been proved to be an useful alternative to overcome the
drawbacks of traditional SPE sorbents and immunosorbents (6,7,8). This adsorptive polymers
are
prepared by a molecular imprinting technique, in which a functional monomer is polymerized with a
cross-linker in the presence of a template molecule. The removal of the template molecule from the
resulting polymer leaves molecularly imprinted complementary binding sites for the template molecule
wich will be mainly retained by specific interactions with functional groups such as hydrogen bonds,
dipole-dipole forces and ion exchange interactions (9). MIPs are therefore used as selective sorbents in
the SPE extraction of target analytes in complex matrices.
On the other hand, it has been shown that organic compounds tend to bind to particulate organic matter
(POM) and dissolved organic matter (DOM) present in environmental water, making the organic
compounds less prone to partitioning with biota (10,11,12,13). In contrast to adsorptive polymers such as
the MIPs, absorptive polymers have a glass transition temperature (Tg) below the temperature of
employment, which causes the absorptive material to operate as a homogenous, non-porous liquid in
which the analytes will be retained by dissolution rather than by specific interactions with the surface of
the polymer (14). This feature allows the absorptive material to equilibrate with the surrounding medium
without reaching saturation (15), in contrast to adsorptive polymers, commonly used in SPE.
2. Extraction and purification of environmental samples.
2.1. Sulfonamides in natural water samples
2.1.1 Offline SPE
The occurrence of 9 sulfonamides and one acetylated metabolite in different environmental waters (waste
water, surface water, ground water) and mineral water, was assessed through three different SPE
approaches and subsequent analysis with LC-QqLIT-MS, using two different adsorptive materials: Oasis
5
HLB cartridges of 200 mg (60 mL) and Oasis MCX cartridges of 150 mg (60 mL) from Waters ( Milford,
MA, USA), and a tandem between both of them. The analytical protocol is detailed in Annex A.
2.1.2 Online SPE
The presence of 19 selected sulfonamides, including one acetylated metabolite, in ground water samples
taken from surveillance and operational monitoring networks located in two ground water bodies in
Catalonia was studied. A new analytical methodology based on online SPE–LC–MS/MS was used.
Briefly, the on-line preconcentration of samples, aqueous standards and operational blanks was performed
using an
automated on-line SPE sample processor (Prospekt-2 TM (Spark Holland, Emmen, The
Netherlands). 40 mL of ground water samples were extracted using Oasis HLB cartridges fromWaters
(Milford, MA, USA). The analytical protocol is detailed in Annex A.
2.2. Triazines and triazine metabolites in soils and natural waters
The presence of nine triazines (chlorotriazines), including three dealkylated metabolites, was studied
through the use of MIP4SPE Triazine10 MIP cartridges (terbuthylazine-imprinted polymer, 25 mg, 10
mL) for the isolation and purification of both water and soil samples. These cartridges were kindly
provided by MIP Technologies AB (Lund, Sweden).
The analysis was carried out by LC-QqLIT MS. To avoid potential contamination from residues of the
MIP template molecule, the analysis of terbuthylazine was carried out through the determination of its
metabolite deethylterbutylazine. The MIP used was commercially available, which considerably
simplifies the pre-treatment step and guaranties the reproducibility between batches of the cartridges to be
used. To assess the selectivity of the MISPE procedure developed for the extraction of triazines and
structurally related compounds, two OPPs, chlorpyrifos and diazinon, were also included as target
analytes. The analytical protocol is detailed in Annex B.
2.3. Polar organic compounds in waste waters
A novel plastic material Poly(ethylene-co-vinyl acetate-co-carbon monoxide) (PEVAC, Figure 1) was
developed to test its ability to assess the bioavailable concentration of a variety of POCs in laboratory
experiments and in wastewater from a sewage treatment plant. Imidacloprid, carbendazim, atrazin,
diazinon, chlorpyrifos, carbamazepine and metoprolol were the compounds selected for the
6
bioavailability experiment. Paracetamol (acetaminophen) , oxazepam , carbamazepine and metoprolol
were the pharmaceuticals selected for the waste water experiments.
The results were compared with existing SPE techniques, using different SPE materials (Oasis MCX and
MAX SPE cartridges, and ENV+ cartridges).
FIGURE 1. The PEVAC polymer (Tg=-32ºC), with the weight ratio of x=66 Wt%, y=24 Wt% and z=10 Wt%
The analytical protocol is detailed in Annex C.
1. RESULTS
3.1. Sulfonamides
3.1.1. Offline SPE-LC-QqLIT-MS
Detailed information regarding the method optimization, recovery values obtained for each adsorptive
material etc can be found in reference 16. For both offline and online SPE methodologies, quantification
was performed based on peak areas and using an internal standard calibration method, crucial to correct
potential matrix effects.
The highest sulfonamide concentrations were those corresponding to surface water from the Llobregat
River, likely due to the many agricultural areas located upstream of the sampling point, as well as to the
proximity of Barcelona city. Figure 2 shows the chromatograms corresponding to one of the three
replicates of the surface waters taken during the first campaign (C1).
7
Figure 2. Total ion current and ion chromatograms corresponding to the SPE extraction and LC–QqLIT–MS
analysis of a natural surface water sample from the Llobregat River.
8
As expected, the lowest concentrations corresponded to the bottled mineral water and the ground water of
Barcelona, in which nearly all the sulfonamides detected were below the LOD (see Annex A).
Sulfadimethoxine was the most frequently SAs residue detected; being present in all water samples, even
in the mineral and groundwater (16). Sulfamethoxazole was also often detected in natural waters and at
high concentrations (maximum 1488 ng L−1), likely because it is used in both human (anti-acne agent)
and veterinary medicine (antibiotic and growth promoter).
Despite that, the highest concentration found in the natural water samples corresponded to sulfapyridine,
which was found to be present in river water up to 12,000 ng L−1.
3.1.2. Online SPE-LC-QqLIT-MS
Ground water samples were taken in spring 2008 in two ground water bodies in Catalonia. A total of 65
wells were sampled, including monitoring wells and natural springs. Ground water was sampled from
depths ranging from 3 to 206 m. Further information regarding the optimization of the online SPE
procedure can be found in reference 17.
Table 1. Descriptive univariate statistics of the data. SD,: standard deviation; *:municipalities and ground
water bodies correspond to the location of the maximum concentration value detected for each sulfonamide.
Contaminants such as sulfonamides and nitrates, highly soluble in water, may reach the water table and
be transported by the slowly moving ground water, widening its presence through very extensive ground
water systems. Since sulfonamides are related to livestock veterinary practices, they could be used as a
specific indicator of manure contamination (18). Nitrogen in its different species (organic, ammonium,
nitrite and nitrate) is a major constituent of manure, and, similarly to sulfonamides, is very soluble in
water. Increased concentrations of nitrate that result from both nitrification of ammonium or direct
introduction from mineral fertilizers are commonly present in both ground water and surface water
associated with ammended agricultural lands. As nitrates and sulfonamides detected in ground waters
9
may share a common origin (extensive and intensive cattle farming and manure application in crop
lands), the possibility of establishing a correlation between both parameters was worth to consider. The
depth of the sampling site and total organic carbon (TOC) were also considered to play an important role
in the concentration distribution of both nitrates and sulfonamide concentrations. With this aim, pair wise
correlations between nitrates concentration, depth of the well and sulfonamides concentration were
estimated and are given in Table 2.
Regarding nitrates and sulfonamides, a clear correlation between the occurrence of both could not be
established from the data obtained. Sulfadiazine and sulfamethizole are the two sulfonamides with the
highest correlation coefficients regarding nitrates concentration (0.37 and 0.33, respectively). However
both sulfadiazine and sulfamethizole were the compounds with the lowest frequencies of detection in both
ground water bodies ( see Table 1). Therefore, the relationship between both nitrates and these antibiotics
should be investigated in further detail.
Table 2. Relationship between depth, nitrate and sulphonamide concentrations, expressed as pairwise
correlation coefficients. The variables are the concentrations of the corresponding sulfonamides (see
reference 18). Higher pairwise correlations are marked in bold.
3.2. Triazines
Detailed information regarding the recovery studies and results for the target triazines can be found in
reference 19. Results listed in Table 3 showed that simazine was the most ubiquous analyte in ground
water and at the highest concentrations, with values up to 105.11 ng L-1. This value was above 100 ng L-1,
the boundary value established in the Directive 98/83/EC for the total amount of pesticides in drinking
water. The other analytes were less frequently detected and at much lower concentrations. For instance,
atrazine was found in three out of the seven wells sampled and DEA and DET in two of them. Propazine
10
and prometon were detected only in one ground water sample, and sebuthylazine and cyanazine were not
detected.
Three soil samples from different agricultural areas fertilized with sludge from a near wastewater
treatment plant, and one sample of that sludge were analyzed. The highest concentrations were found
Figure 3: Comparison between the chromatograms of a sample extracted using the triazine MIP and a nonimprinted polymer extracting material in a HPLC grade water sample spiked at 2.5 ng L-1 .
in the most recently sludge-fertilized soils, S1 and S3 (see Table 4). Sebuthylazine was the analyte
detected at the highest concentrations, with values in the range 0.46-2.26 ng g-1. Propazine was detected at
the lowest levels, from 0.05 to 0.11 ng g-1. Sludge showed the lowest concentrations of triazines (0.010.28 ng g-1); this evidenced that the contribution to the pesticide load from the sludge was not comparable
to the input by means of spreading of pesticides over the fields. The fact that more recently fertilized soils
contained higher concentration of pesticides might be due to the combined actions of spreading and
simultaneous sludge application at the same stage of the crop.
11
Table 3: Concentrations (ng L-1) in the ground water samples investigated. - : below the method limit of
detection.
Table 4: Concentrations (ng kg-1) in the sludge and the sludge fertilized soils investigated.
3.3. Polar organic contaminants
3.3.1. Evaluation of fulvic acid enrichment on the PEVAC sampler and on a polar resin based SPE
cartridge.
Conventional SPE-methods are designed to quantify the total amount of an analyte from the sample
matrix. The SPE-technique has two major drawbacks when assessing the freely dissolved fraction of
POCs. First, the exhaustive extraction approach may disturb the initial distribution of the analytes
between water, DOM and POM in the sample (15). Secondly, polar resin based cartridges for efficient
recovery of POCs will also extract DOM (2), making it impossible to distinguish the DOM-bound
fraction of an analyte from the freely dissolved fraction.
The chromatogram obtained from the SPE extraction of FA rich water in this work presented the
accumulated FA as a broad hump extending from 1.5 to 4.0 minutes (Figure 4a). The summed
massspectrum of the hump further confirmed the presence of FA in the extract. Thus, it is impossible to
differentiate the contribution to a chromatographic peak of a chemical, which can originate both from the
freely dissolved concentration (CFree) and from the FA bound concentration. The removal of humic
substances in an extraction method is therefore crucial when it comes to assessing the CFree of a substance
present in aquatic environments. A summed mass spectrum over the same area and retention time as the
hump in the SPE chromatogram was was also constructed from the chromatogram obtained from the
PEVAC sampler. The mass spectrum revealed the absence of FA in the PEVAC extract (Figure 4b),
which indicates that measurements performed with the PEVAC sampler will represent the truly dissolved
12
concentration of contaminants in FA rich water given that negligible depletion is achieved. The same
result has earlier been reported, where the SPE method gave a humic associated hump in the beginning of
the chromatogram which disappeared when supported liquid membrane (SLM) extraction was applied, a
precursor to the ESTM sampler, on a HS rich river water (20).
Figure 4. The ability to enrich fulvic acid on a; a. polar resin based SPE cartridge, b. on a PEVAC sampler
3.3.2. Influence of fulvic acid or suspended sediment on the freely dissolved concentration of POCs
The knowledge that measurements performed with the PEVAC sampler assess mainly the CFree of
chemicals in environmental water, make it possible to evaluate the abundance of the seven POCs to DOM
and POM.
Figure 5a, show that only Metoprolol out of the seven POCs presented a decrease in CFree as the amount
of FA in the water increased. The result can seem confusing when earlier findings has reported the
binding between contaminants and DOM to rely on hydrophobic interaction (10), and metoprolol is one
of the less hydrophobic compound out of the seven. However, Lützhøft and co-workers et al 2000 (22)
13
investigated ionisable compounds ability to bind to DOM in water at different pH. They found that the
ability for a compound to bind to DOM is highly dependent on pH and that organic cations can
electrostatically interact with anionic sites present in the FA structure at pH >4. Metoprolol is the only
compound of the investigated seven substances having a pKa sufficiently low to make it mainly
protonated and positively charge at pH 7. Thus ionic bonding between Metoprolol and DOM is a
plausible explanation for the drastic decrease in CFree as the FA concentration is increased, since the FA
used in the experiment becomes deprotonated at pH > 3.8. However, a decrease in CFree based on just
hydrophobic interaction with DOM is not likely for a compound with Log Kow< 5 (11), why no visible
decrease is seen for the other six compounds as the FA concentration increased.
Figure 5. The freely dissolved concentration (CFree) of the seven selected compounds as a function of; a. the
amount of fulvic acid in water, b. the amount of sediment in water
Important to note in the former experiment is that the water concentration of all the POCs were constant
throughout the experiment independently of the added amounts of FA indicating that negligible depletion
was achieved, except for diazinon and chlorpyrifos, where none of the fifteen samples proved to have
detectable levels of the two contaminants (Data not presented). Diazinon and chlorpyrifos were probably
depleted from the water phase due to the high enrichment factor in the PEVAC film, which in turn makes
the estimation of them unreliable. However, if the amount of a chemical bound to DOM or POM is much
greater than the amount finally absorbed by the sampler a temporary depletion of the chemical in the
water phase is allowed, because when steady state in the sample is finally stabilized, net-desorption from
the DOM or the POM will re-establish the initial CFree in the water phase. In these cases it is the DOM or
the POM and not the water that represents the matrix that should not be depleted (21).
When the FA, representing the DOM, was replaced with sediment, the CFree of all the seven POCs was
more or less affected by the increasing amount of sediment in the water sample (Figure 5b). Chlorpyrifos
and diazinon showed a significant decrease in CFree as the sediment concentration increased, probably due
to hydrophobic interaction with fattier and higher amount of TOC in the sediment experiment than in the
DOM experiment. The decrease in CFree for metoprolol was even more pronounced when the sediment
14
concentration was elevated compared to the increase of DOM in the earlier experiment. The results
suggest that sediment contain charged groups other than the deprotonated carboxylic groups represented
in the FA. Solid particles and minerals often have negative sites exposed on their surface that will interact
with positively charged organic and inorganic ions (25, 26). The excess of negatively charged sites on
particulate surfaces may explain why the CFree of carbendazim drastically dropped when the sediment
concentration increased in comparison with the previous DOM experiment. The major part of
carbendazim is represented in its neutral form at pH 7, but since it has a pKb-value of 6.09, around 10%
will be protonated and positively charged at pH 7. The irreversible interaction between the positively
charged carbendazim and the excess of negatively charged sites in the sediment will deplete the initial
ionic fraction of carbendazim from the aqueous phase. The remaining neutral fraction of carbendazim in
the aqueous phase will then gradually be rearrange into more positively charged ions interacting with the
sediment until the freely dissolved fraction of carbendazim is depleted or until the surface of the sediment
is saturated.
The water concentrations of the seven POCs throughout the sediment experiment were not possible to
establish due to a high amounts of dispersed particulate matter in the water phase.
The conclusion is that total extraction is a sufficient method to estimate the freely dissolved concentration
of uncharged POCs in natural water. Furthermore, the work illustrated that the PEVAC polymer can be an
alternative material to silicone and PA when it comes to estimating the bioavailable concentration of
POCs in environmental matrices.
3.3.3. Sampling of wastewater from a sewage treatment plant.
For three of the four pharmaceuticals utilized in the experiment the two sampling techniques showed no
significant difference in estimating the freely dissolved concentration of the compounds through out the
sewage treatment plant (Figure 6 - 8). The result showed that a total organic carbon (TOC) with an
average of 150 mg/L in the influent water and 9.6 mg/L in the effluent water is too low to, other than
marginally, influence the freely dissolved concentration (CFree) of the pharmaceuticals. The fourth
compound metoprolol, which is the only compound out of the four pharmaceuticals that is cationic under
the prevailing pH condition, revealed a decrease of 43% in freely dissolved concentration (C Free) in the
influent water when quantified utilizing the PEVAC sampler and compared to the result from the SPE
extraction (Figure 9).
15
Figure 6. A comparison of paracetamol concentrations in treated and untreated waste water established with SPE and
PEVAC extraction techniques.
Figure 7. A comparison of Carbamazepine concentrations in treated and untreated wastewater established with SPE and
PEVAC extraction techniques.
The result corresponds reasonably well with the previous laboratory experiment, regarding the influence
of FA on the freely dissolved concentration of metoprolol (Figure 5a). However, the activated sludge and
sand filter processed effluent water showed an increase of the bioavailable concentration of metoprolol of
366% when the PEVAC sampler was utilized compared to the result from the SPE extraction (Figure 9).
The enhanced bioavailable concentration of metoprolol in the effluent water appears to contradict
previous results. However, under the laboratory investigation the pH was constant (Figure 5a), which was
not the case in the sampling of the wastewater. An increase of pH in the effluent water would severely
change the distribution between the ionic and non-ionic fraction of metoprolol in the water phase and thus
the absorbent/water partition coefficient (KAbs/W) (See deliverable 2.2.17).
16
Figure 8. A comparison of oxazepam concentrations in treated and untreated wastewater established with SPE and PEVAC
extraction techniques.
If only the non-ionic fraction of metoprolol is assumed to partition with the absorbent of the PEVAC
sampler an increase of pH from 7.8 in the influent water to 8.4 in the effluent water would lead to a 355%
increase in non-ionic fraction of metoprolol and thus a 355% increase of KAbs/W . However, the Kabs/w
used for the sewage water was the one that was determined at pH 7, thus the elevated concentration of
metoprolol should be regarded as an artefact. This demonstrates the importance of using the appropriate
Kabs/w, especially for charged species, when applying this sampling technique to real samples.
Figure 9. A comparison of metoprolol concentrations in treated and untreated wastewater established with SPE and
PEVAC extraction techniques.
The analysis of the waters showed that activated sludge, sand filtration in combination with ozone
treatment was the most effective approach to eliminate pharmaceuticals from the wastewater (Figures 6 to
9). Activated sludge and sand filtration alone (effluent) as well as membrane bioreactor (MBR) was not
sufficient treatment to significantly reduce the freely dissolved concentration (C Free) of the selected
17
compound in the wastewater, except in case of paracetamol, for which a drastic decrease was observed
with all applied treatment techniques.
CONCLUSIONS
First, the two developed methods (following offline and online SPE strategies) were proved to be
powerful tools for the analysis of sulfonamides and N4-acetylated metabolites in complex natural waters,
allowing determining concentrations down to the ng L−1 level in all kind of waters (wastewater,
groundwater, surface and bottled mineral water). Results on groundwater and mineral bottled water
constitute the first data reported on levels of SAs in these types of waters. On the basis of these outcomes,
the N4-acetylated metabolites of sulfonamide antibiotics should be considered in environmental residue
analysis to avoid the underestimation of the elimination rates of SAs during wastewater treatments. In the
offline SPE study, N4-acetylsulfamethazine was detected in all the samples at concentration up to 316 ng
L−1 ( concentration detected in surface water).
Regarding the online SPE methodology applied, 18 out of the 19 target sulfonamides were detected in the
two ground water bodies from Catalonia studied, being sulfadimethoxine and sulfamethazine, commonly
used in veterinary practices, those occurring more frequently. It should be highlighted again the high
frequency of detection for the acetylated metabolite N4-acetylsulfamethazine, comparable to the highest
frequencies aforementioned. The need for the inclusion of this and other metabolism products in future
monitoring studies is unquestionable. Despite the peak concentration values detected in different
sampling sites, the average detected concentrations of sulfonamides are generally below 50 ng/L.
Sensitivity is therefore one of the most critical parameters in order to obtain unequivocal and reliable
determination for he compounds investigated. When performing on-line SPE analysis, its fully
automation and the minimum sample manipulation requirements permits the enhancement of sensitivity,
as the whole sample volume (40 mL) gets to the chromatographic system instead of a final reconstituted
extract as in off-line procedures, where usually volumes of 200 mL or bigger are reduced to
approximately 0.5 mL and only around 10 µL will be injected in the mass analyzer. Despite the low
sample volumes required in on-line procedures, it has been proved that sensitivity is not affected but, on
the contrary, improved considerably, with limits of detection down to the pg/L level. Besides, LC–
MS/MS allows for an unequivocal identification of the target sulfonamides. From the results obtained, no
strong correlation between sulfonamides and nitrates concentrations could be established. Whereas
nitrates in ground water are originated from fertilizers of both animal and mineral origin, sulfonamides
could be specifically considered as potential indicators of pollution from animal origin. For this reason,
and because data on nitrates is historically richer and more consistent, the presence of sulfonamides in
ground water matrices should be investigated in further detail in order to propose these substances as
reference points to indicate pollution from animal farm and agriculture practices.
18
Regarding triazines and the use of MIPS, the outcomes obtained evidenced the excellent affinity and
selectivity provided by the Triazine 10 MIP used in the extraction and purification of complex
environmental samples containing triazines and structurally related compounds (see Annex B) .The use of
the LC-MS/MS developed method afforded high sensitivity (LODs in the low ng L-1and ng g-1) and
allowed an unequivocal identification (in SRM mode) of the triazines investigated. The QqLIT hybrid
mass analyzer was the most suitable tool for detection and determination of trace levels of triazines in
complex environmental samples, especially in connection with MISPE isolation and purification.
The levels of the target pesticides found in the natural water analyzed were below the limit established by
the European Legislation for drinking water, except for one ground water sample, where concentrations
detected for simazine and atrazine were 105.11 ng L -1 and 35.37 ng L-1 respectively. For some of the
compounds studied (DIA, DEA, atrazine, prometon, cyanazine), concentrations found in water were
similar to the values detected in soils (0.01-2.26 ng g-1), but the majority were below 0.50 ng g-1. Pesticide
concentrations in the sludge applied as fertilizer were generally low, which indicated that the contribution
of the sludge to the total amount of pesticides in the sludge-fertilized agricultural soils was not as relevant
as the direct pesticide application to fields.
Freely dissolved concentrations determined from equilibrium sampling of uncharged polar organic
compounds using the novel sampler compared well with the total concentration measurements obtained
with the serial SPE method. The results further indicate that the freely dissolved concentrations are lower,
although not drastically lower, than the total concentrations. For species that are mainly ionic, we
conclude that in the case that distribution coefficients are lacking for the prevailing pH, in situ
equilibrium sampling is not recommendable. Instead, equilibrium sampling and quantitation using grab
samples in combination with the use of either isotopically labelled surrogate standards or, alternatively, a
standard addition approach is recommended. Furthermore, biological treatment of municipal waste water
in combination with sand filtration and ozone treatment showed promising results regarding removal of
pharmaceutical residues.
19
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(9) Filippov, O. A.; Tikhomirova, T. I.; Tsizin, G. I.; Zolotov, Y. A. Dynamic preconcentration of
organic substances on nonpolar adsorbents. J. Anal. Chem. 2003, 58, 398-422.
(10) Traina, S. J.; Mcavoy, D. C.; Versteeg, D. J. Association of linear alkylbenzenesulfonates with
dissolved humic substances and its effect on bioavailability. Environ. Sci. Thchnol. 1996, 30,
1300-1309.
(11) Chin, Y. P.; Weber, W. J., Jr.; Eadle, B. J. Estimating the effects of dispersed organic polymer on
the sorption of contaminants by natural solids. 2. Sorption in the presence of humic and other
natural macromolecules. Environ. Sci. Thchnol. 1990, 24, 837-842.
(12) Day, K. E. Effects of dissolved organic carbon on accumulation and acute toxicity of fenvalerate,
deltamethrin and cyhalothrin to Daphnia magna (straus). Environ. Toxicol. Chem. 1990, 10, 91101.
20
(13) Boa, M. L.; Dai, S. G.; Pantani, F. Effect of dissolved humic material on the toxiciy of tributyltin
chloride and triphenyltin chloride to Daphnia magna. Environ. Contam. Toxicol. 1997, 59, 671676.
(14) Baltussen, E.; Cramers, C. A.; Sandra, P. J. F. Sorptive sample preparation – a review. Anal.
Bioanal. Chem. 2002, 373, 3-22.
(15) Mayer, P.; Tolls, J.; Hermens, J. L. M.; Mackay, D. Equilibrium sampling devices. Environ. Sci.
Technol. 2003, 37, 184A-191A.
(16)García Galán, M.J., Díaz-Cruz, S., Barceló, D. Highly sensitive simultaneous determination of
sulfonamide antibiotics and one metabolite in environmental waters by liquid chromatographyquadrupole linear ion trap-mass spectrometry. J. Chrom. A.2008, 1193, 50-59.
(17)García Galán, M.J., Díaz-Cruz, S., Barceló, D. Determination of 19 sulfonamides in
environmental waters by automated on-line solid phase extraction-liquid chromatography-tandem
mass spectrometry (SPE-LC-MS/MS) (submitted)
(18)García Galán, M.J., Garrido, T., Fraile, J., Ginebreda, A., Díaz-Cruz, S., Barceló, D.
Simultaneous occurrence of nitrates and sulfonamide antibiotics in two ground wáter bodies of
Catalonia (Spain). J. Hydrology, 2009 (accepted).
(19)García Galán, M.J., Díaz-Cruz, S, Barceló, D. Determination of triazines and their metabolitesin
environmental samples using molecularly imprinted polymer extraction, pressurized liquid
extraction and LC-tandem mass spectrometry. J. Hydrol., 2009, (accepted).
(20) Megersa, N.; Solomon, T.; Jönsson, J. Å. Supported liquid membrane extraction for sampling
work-up and preconcentration of methoxy-s-triazine herbicides in a flow system. J. chromatogr.
A. 1990, 830, 203-210.
(21) Heringa, M. B.; Hermens, J. L. M. Measurements of free concentrations using negligible
depletion-solid phase microextraction (nd-SPME). Trends in Analytical Chemistry. 2003, 22,
575-587.
(22) Lützhøft, H. C. H.; Vaes, W. H. J.; Freidig, A. P.; Halling-Sørensen, B.; Hermens, J. L. M.
Influence of pH and other modifying factors on the distribution behaviour of 4-quinolones to
solid phases and humic acids studied by “Negligible-Deplition” SPME-HPLC. Environ. Sci.
Technol. 2000, 34, 4989-4994.
(23) Reichenberg, F.; Mayer, P. Two complementary sides of bioavailability: accessibility and
chemical activity of organic contaminants in sediments and soils. Environ. Toxicol. Chem. 2005,
25, 1239-1245.
(24) Semple, K. T.; Doick, K. J.; Jones, K. C.; Burauel, P.; Craven, A.; Harms, H. Defining
bioavailability and bioaccessibility of contaminated soil and sediment is complicated. Environ.
Sci. Technol. 2004, 38, 228A-231A.
21
(25) Schwarzenbach, R. P.; Gschwend, P. M.; Imboden, D. M. Environmental Organic Chemistry.
Second ed. John Wiley & Sons Inc. Hoboken, New York, 2003.
(26) Haderlein, S. B.; Schwarzenbach, R. P. Adsorption of substituted nitrobenzenes and nitrophenols
to mineral surfaces. Environ. Sci. Technol. 1993, 27, 316-326.
ANNEX A
ANALYTICAL PROTOCOLS FOR THE ANALYSIS OF SELECTED SULFONAMIDES IN
DIFFERENT NATURAL WATERS
1. CHEMICALS





Organic solvents: Methanol, Acetone, Acetonitrile, all HPLC grade
Water HPLC grade
Formic acid.
Sulfonamides:
o Offline SPE procedure: sulfadiazine, sulfadimethoxine, sulfamethazine,
sulfamethizole, sulfamethoxazole, sulfamethoxypyridazine, sulfapyridine,
sulfisoxazole, sulfathiazole and N4-acetylsulfamethazine.
o Online SPE procedure: the sulfonamides aforementioned plus sulfadimidin,
sulfamerazine, sulfacetamide, sulfabenzamide, succinyl-sulfathiazole,
sulfaguanidine, sulfanitran, sulfaquinoxaline and sulfadoxine.
Oasis HLB and MCX SPEcartridges (Waters).
2. SAMPLE COLLECTION AND PRETREATMENT
Surface water samples were taken from the Segre River (Lleida, Spain) and Anoia River (Barcelona) in
June 2006. Samples from the wastewater treatment plant (WWTP) in Lleida were also sampled in June
2006. In May 2007 groundwater from the city of Barcelona was also sampled. Bottled mineral water was
purchased from a supermarket. Environmental water samples, around 1 L, were shipped to our laboratory
under cool conditions in amber glass bottles pre-rinsed with HPLC grade water in a portable refrigerator,
readily vacuum filtered through 2.7 µm glass fiber filters followed by 0.45 µm nylon membrane filters
(Whatman, Maidstone, UK) and stored at 4 ◦C until analysis (less than 2 days). Under such conditions
ny antibiotic activity in the samples was kept to the minimum. In the offline SPE procedure, three sample
replicates were taken for each water matrix.
Ground water samples for the online SPE procedure were taken in spring 2008 in two ground water
bodies in Catalonia: Plana de Vic and La Selva. Water samples were collected in amber polyethylene
terephthalate (PET) bottles and transported to the laboratory under cooled conditions (4 ºC). Once there,
samples were filtered through 0.45 lm Nylon filters to eliminate suspended solid matter and then stored at
22
4 ºC in the dark until analysis which was always carried out within 48 h of collection to avoid microbial
degradation.
3. SOLID PHASE EXTRACTION
The analytical protocol followed for the SPE extraction was adapted from a methodology previously
developed and applied by the authors in the purification of pressurized liquid extracts of sludge samples
for the analysis of SAs. Figure 1 shows the three protocols followed. The method using the HLB
cartridges was shown to be the most efficient. Under the selected procedure, 400mL of water samples
(200mL for waste water) were loaded at 3mL min−1 onto Oasis HLB cartridges preconditioned with 3mL
of MeOH in 50mM HCOOH, followed by 3mL of acetone in 50mMHCOOH and 2mL of HPLC grade
water at neutral pH with a 5% ofMeOH. After the sample extraction, the cartridges were rinsed with 2mL
of HPLC grade water at neutral pH with a 5% of MeOH, to remove potential interferences. Finally,
cartridges were dried under vacuum (around 30 min.) and then eluted with 6mL of MeOH in 50mM
HCOOH plus 6mL of acetone in 50mM HCOOH at 3mLmin−1. The resulting eluates were brought to
dryness under a gentle N2 stream and reconstituted with 400µL of MeOH containing d4-sulfathiazole as
internal standard, for further analysis.
Regarding the online SPE protocol, the procedure had been optimized in a previous work (REFERENCE)
the on-line preconcentration of samples, aqueous standards and operational blanks was performed using
an automated on-line SPE sample processor Prospekt-2 TM (Spark Holland, Emmen, The Netherlands).
The online Oasis HLB cartridges were conditioned with 1 mL of a mixture of methanol/acetone (1:1 v/v)
at a rate of 1 ml min-1 and equilibrated with 1mL of water.40 mL of the ground water samples were
loaded onto the cartridges, which were washed afterwards with 1 mL of water at a flow rate of 1 mL min1 to improve the complete transfer of the sample and remove interferences. Cartridges are then transferred
to the elution clamp and the analytes are eluted directly onto the LC column by the HPLC and the
gradient solvents.
23
Figure 1. Different solid-phase extraction strategies followed: (a) HLB cartridges, (b) MCX cartridges and (c)
tandem extraction (HLB and MCX).
24
4. ANALYTICAL DETERMINATION

Instrument: HP 1100 chromatograph (Agilent Technologies, Palo A, CA, USA) coupled to a
4000 QTRAP mass spectrometer (Applied Biosystems, Foster City, USA) equipped with a
turbospray electrospray source.

LC column: Atlantis C18 (Waters, 150 mm x 2.1 mm, 3 µm) preceded by a guard column
with the same packing material.

Ionization: positive ionization mode (PI).

Gradient elution: mobile phase of HPLC grade water and acetonitrile, both slightly acidified
with 1mM formic acid. The gradient is programmed from 25% to 100% of the organic phase
in 11 minutes, held for two minutes and returned to the initial conditions in 10 minutes.

Flow rate: 0.2 mL min-1

Injection volume: (offline procedure): 20 µL.
5. METHODS VALIDATION

Calibration curves: internal standard calibration based on peak areas (concentrations ranging
between 0.05 and 1000 ng L−1)

Method recoveries (offline SPE method): please,see reference 16.

Limits of detection and quantification: please, see Table 1 for the offline SPE procedure and
Table 2 for the online SPE procedure.
Table 1. Recovery rates obtained for the Oasis HLB approach, precision, expressed as relative standard
deviation RSD (%), and method limits of detection and quantification for the different water matrices
studied. Recovery values below 50% or over 150% were neglected (–).
25
Table 2. Performance of the on-line SPE–LC-QqLIT–MS method applied. r2, correlation coefficient; LOD,
method limit of detection; LOQ, method limit of quantification and RSD, relative standard deviation (%).
ANNEX B
ANALYTICAL PROTOCOL FOR THE ANALYSIS OF TRIAZINES AND TRIAZINE
METABOLITES IN WATER AND SOIL SAMPLES.
1. CHEMICALS:

Organic solvents: methanol, acetonitrile, dichloromethane (DCM), all HPLC grade

Water HPLC grade

Pesticides: atrazine, simazine, cyanazine, sebuthylazine, deisopropylatrazine (DIA),
deethylatrazine (DEA) and deethylterbuthylazine (DET), prometon, propazine, chlorpyrifos
and diazinon d5-atrazine (100 μg mL-1, acetone) and the surrogate standard d5deisopropylatrazine

Hydromatrix (Varian)

Oasis HLB cartridges (Waters)

MIP4SPE Triazine10 MIP cartridges (terbuthylazine-imprinted polymer, 25 mg, 10 mL)
26
2. SAMPLE COLLECTION AND PRETREATMENT
Ground water, tap water and HPLC grade water were the selected samples for this study. Tap water from
the city of Barcelona (Catalonia, Spain) and ground water from seven wells along the province of
Barcelona were taken in March 2007. Ground water samples (1 L each) were shipped to our laboratory
under cool conditions (4ºC) in amber glass bottles pre-rinsed with HPLC grade water in a portable
refrigerator, readily vacuum filtered through 2.7 m glass fiber filters followed by 0.45 m nylon
membrane filters and stored at 4 ºC until analysis (less than two days). Tap water samples were taken the
day before the analysis and were not filtered. Three sample replicates were taken in all cases.
Soil samples were collected at the end of October 2005 at selected sites from different intensive
horticulture areas in the proximity of Barcelona (Catalonia, northeast of Spain). The soil was calcareous
type (c.a. 30% sand, 25% silt and 45% clay) with a pH 7-9 and a low to moderate content of organic
carbon (0.9-3.5%). Representative sampling of the top soil (0-20 cm) was performed with an Auger
sampler (Eijkelkamp, Giesbeek, The Netherlands) at a rate of five subsamples per hectare. Individual
subsamples were shipped in aluminum foil packets to our laboratory under cool conditions (4º C) in a
portable refrigerator. Subsamples corresponding to the same agricultural field were pooled, and then
readily frozen at –20 ºC upon arrival. Samples were afterwards lyophilized, homogenized, sieved through
a stainless steel 0.2 mm sieve and stored at –20 ºC in sealed containers until analysis.
Subsamples of soil were taken from a site in the same area where sludge was not applied to the soil. This
pooled sample was used as control blank soil. Samples of the sewage sludge applied in the soil
fertilization process were also taken.
3. CLEAN UP
Aliquots of 10 ml of the water samples were passed through the MIP cartridges previously conditioned
with 1 mL of MeOH, 1 mL of HPLC grade water and 1 mL of 50 mM NH4H2PO4 solution at pH 3, at a
27
flow rate of 3 ml min-1 using a 12-fold vacuum extraction box (J.T. Baker, Phillipsburg, NY, USA). To
avoid interferences and eliminate potential matrix effects, the cartridges were then washed with 1 mL 0.1
M HCl and 1 mL HPLC grade water. After that and in order to boost selective interactions between the
MIP and the target analytes, 1.5 mL of DCM were loaded onto the cartridges. Vacuum (20 min) was
applied before adding the DCM, since water and DCM are not miscible, and 2 min more after adding the
DCM, to remove the remains of the solvent. As DCM is a weakly polar and aprotic solvent, it favors the
interactions between the analytes and the binding sites, and simultaneously removes co-extracted
substances from the MIPs. Finally, the cartridges were eluted with 3 mL MeOH at a slower flow rate than
before (2 ml min-1). Since it is a very polar and protic solvent, it disrupts the hydrogen bonds between the
polymer and the compounds, allowing the elution. The resulting eluates were evaporated under a gentle
N2 stream at 25 ºC in a Turbo Vap LV evaporator (Zymark, Hopkinton, MA, USA), then reconstituted
with 0.5 mL of LC mobile phase (methanol:water) at the initial conditions (10:90, v:v).
The extraction of the solid samples was carried out by pressurized liquid extraction using an accelerated
solvent extractor ASE 200 (Dionex). 2 g of the lyophilized and sieved soil/sludge samples were mixed
with Hydromatrix (previously ultra-sonicated with a mixture of acetone:MeOH (50:50, v:v) three times
during 10 min. each one) into an extraction cell sealed at both ends with glass-fiber filters. Extraction was
performed with a mixture of MeOH:acetone (50:50, v:v) under the following experimental conditions:
pressure, 1500 psi; temperature, 65 ºC; preheat time, 5 min.; static time, 5 min.; extraction time, 3 min.;
flush volume, 60%; purge time, 60 s., and 3 extraction cycles.
The PLE extracts were evaporated under a gentle N2 stream and reconstituted with a mixture of 1 mL
MeOH and 19 mL HPLC grade water. 10 mL aliquots were then purified following the MISPE procedure
previously optimized for the extraction of triazines from the water samples.
4. ANALYTICAL DETERMINATION

Instrument: HP 1100 chromatograph (Agilent Technologies, Palo A, CA, USA) coupled to a
4000 QTRAP mass spectrometer (Applied Biosystems, Foster City, USA) equipped with a
turbospray electrospray source.
28

LC column: Purospher STAR RP-18 endcapped (125 mm x 2 mm; 5μm) (Merck, Darmstadt,
Germany) preceded by a guard column with the same packing material.

Ionization: positive ionization mode (PI).

Elution gradient: mobile phase of HPLC grade water and methanol. The gradient progressed
from 10 % to 100 % of the organic phase in 22 min, held there for 3 minutes, returned to the
initial conditions in 5 minutes and, finally, a reequilibration period of 5 min was
programmed.

Gradient flow rate: 0.2 ml min-1

Injection volume: 20µL.
5. METHOD VALIDATION

Calibration curves: internal standard calibration based on peak areas (concentrations ranging
between 0.01 - 40 ng mL- 1).

Method recoveries (offline SPE method): please, see reference 19.

Limits of detection and quantification: please, see reference 19.
ANNEX C. ANALYTICAL PROTOCOL FOR THE ANALYSIS OF POLAR
ORGANIC CONTAMINANTS.
1. CHEMICALS:

Organic solvents: toluene, methanol, acetonitrile, dichloromethane (DCM).

Ohers: acetic acid, formic acid, hydrochloric acid, ammonium bicarbonate, ammonium
hydroxide solution (25%). Poly(ethylene-co-vinyl acetate-co-carbon monoxide) beads.
Nordic Reference Fulvic acid

POCs: Imidacloprid, carbendazim, atrazin, diazinon, chlorpyrifos, carbamazepine, metoprolol
(tartrate
salt),
paracetamol
(acetaminophen)
,
[2H5]Oxazepam
(99%
purity)
,
[2H10]Carbamazepine (98.2 atom% 2H) and [2H3]paracetamol (99.1% atom% 2H) .

ENV+ columns (50mg, 60µm, 3mL) were purchased from Isolute (Hengoed, UK).

Oasis MCX and MAX SPE columns (60 mg, 30 µm, 3 mL)
2. PROCEDURES
29
2.1. PEVAC sampler preparation.
(See Deliverable 2.2.17)
2.2. Evaluation of fulvic acid enrichment on the PEVAC sampler and on a polar resin based SPE
cartridge.
Three 200mg PEVAC samplers were left in a glass-vessel filled with 3.0 litres of tap-water for 2 days
under gentle stirring. The water was spiked with 50mg/L of FA and buffered to pH 7 with ammonium
bicarbonate. After 2 days, when equilibrium was achieved, the PEVAC samplers were removed from the
solutions and extracted with MeOH. The extract was evaporated to dryness under nitrogen at 40°C and
redissolved in 200µl of MeOH: Milli-Q water (1:1).
20ml of the FA rich water solution was filtrated through a SPE cartridge containing 50 mg ENV +, a polar
divinylbenzene-polystyrene (DVB-PS) based resin. The filtrate was extracted from the SPE cartridge
using MeOH. The extract was evaporated to dryness under nitrogen at 40°C and redissolved in 200µl of
MeOH: Milli-Q water (1:1).
2.3.Influence of fulvic acid or suspended sediment on the freely dissolved concentration of
POCs.
Fifteen glass-bottles were filled to the edge with 120ml of tap-water. The water was spiked with
imidacloprid, carbendazim, metoprolol, atrazin, carbamazepine, diazinon and chlorpyrifos. In the DOM
experiment various amounts of FA were added to the bottles, providing concentrations of 0, 5, 10, 50 and
100 mg/L of FA in the bottles. In the POM experiment various amounts of natural sediment, with a total
organic carbon (TOC) content of 2.2%, were added to the bottles, providing concentrations of 0, 5, 10, 50
and 100 g/L of suspended sediment in the bottles. All the samples were buffered to pH 7 with ammonium
bicarbonate. 20mg of dry PEVAC-film were added to each of the fifteen bottles in the two experiments.
The glass-bottles were then rotated for 2 days, which was the pre-determined equilibrium time (See
Deliverable 2.2.17).
After 2 days, the PEVAC samplers were removed from the solutions and extracted with MeOH. The
extract was evaporated to dryness under nitrogen at 40°C and redissolved in 100µl of MeOH. The
remaining concentrations of the seven POCs in the water-sample spiked with FA was established, from
each sampling occasion, by evaporating 1.0ml of water to dryness under nitrogen at 40°C and redissolve
it in 100µl of MeOH
30
2.4.Sampling of wastewater from a sewage treatment plant
24-hour composite samples were collected during four days from Hammarby Sjöstad sewage treatment
plant (STP), Stockholm, Sweden. The influent water was treated by a conventional activated sludge
process, using a sludge residence time of five days, followed by sand filtration. Additionally, a fraction of
the biologically treated effluent water was passed through an ozone treatment step, using 15 g O 3/m3 of
wastewater. In a separate treatment process, influent water was transferred to a membrane bioreactor
(MBR) (Kubota Submerged Membrane Unit), in parallel to the activated sludge treatment. Samples were
collected from the influent stream, from wastewater treated by activated sludge and a final sand filter,
from ozone treated effluent and from the MBR. The samples, collected in plastic bottles, were
immediately frozen and stored at -20° C until further analysis. When thawed, samples collected from four
different days were combined into one pooled sample.
2.5. SPE extraction of wastewater.
The serial cationic-, anionic-exchange SPE method used for determination of pharmaceutical residues in
waste water is described in detail by Lavén et al. (20). In short, wastewater samples were initially passed
through a glass-microfibre filter to remove particulate matter. An aliquot of 50 mL was used for the
analysis of effluent, ozone and MBR treated wastewater, whereas 25 mL of wastewater was sufficient for
the analysis of influent water. Additionally, 25 mL of H2O was added to influent water samples. The
sample was adjusted to pH 2 with HCl (37%), and deuterated surrogate standards of paracetamol,
carbamazepine and oxazepam were added prior to the SPE step. Metoprolol standard was added as a
standard addition to duplicate samples prior to the SPE extraction. The clean-up procedure results in three
fractions containing basic (eg. oxazepam and metoprolol), neutral (e.g.,paracetamol and carbamazepine)
and acidic analytes, respectively.
The SPE fractions were evaporated to dryness under nitrogen at 40°C and redissolved in 500 µL of 20%
acetonitrile, 0.1% formic acid, prior to LC-MS analysis.
2.6. PEVAC extraction of wastewater
Glass-bottles were filled to the edge with 120ml of influent wastewater or effluent, ozone and MBR
treated wastewater. Deuterated surrogate standards of paracetamol, carbamazepine and oxazepam were
added to the bottle prior to sampling. 100mg of PEVAC-film were added to each of the glass-bottles in
the experiments. The bottles were then rotated for 2 days, which was the pre-determined equilibrium time
(See Deliverable 2.2.17). The samples were not pH-adjusted.
31
After 2 days, the PEVAC samplers were removed from the solutions and extracted with MeOH. The
extract was evaporated to dryness under nitrogen at 40°C and redissolved in 100µl of MeOH, 0.1%
formic acid, prior to LC-MS analysis.
3. ANALYTICAL DETERMINATION

Instrument: ACQUITY Ultra performance Liquid Chromatograph (Waters, Milford, USA)
coupled to a QTOF.

LC column: Acquity HSS T3 C18 column (1.7 µm, 2.1x100mm).

Gradient elution: mobile-phase mixture of 95% Milli-Q water and 5% acetonitrile (ACN)
buffered with 10mM acetic acid. The percentage of ACN in the mobile-phase linearly
increased to 95% for 5 minutes and was held at 95% ACN for 2.5 minutes. After a total of 7.5
minutes the percentage of ACN linearly decreased for 0.5 minutes to the initial mobile-phase
composition and remained there to the end of the run for a total of 11 minutes.

Flow-rate: 0.3 ml/min.

Ionization: positive ionization mode (PI).
4. METHOD PERFORMANCE
The recovery and ion suppression of the SPE method were studied by an experimental set-up using the
following samples, in triplicates: Non-spiked wastewater (Non-Spiked), Wastewater spiked prior to
extraction (Pre-Extr), Wastewater spiked after extraction in the reconstitution step (Post-Extr), and
Spiking solution (Sp-Sol) (20). The recovery was calculated from the “apparent” analyte concentration of
samples spiked prior to and after extraction, using equation (2):
Recovery = (CPre-Extr-CNon-Spiked)/(CPost-Extr-CNon-Spiked) x 100
(2)
where C denotes the “apparent concentration”. Since the samples already contained a number of analytes,
the native contribution to the “apparent concentration” of spiked samples was subtracted. In the case of
group 1 analytes, the deuterated analogues were used as spikes. Ion suppression was calculated using the
equation (3):
Ion suppression (%) = (1-(CPost-Extr-CNon-Spiked)/(CSp-Sol)) x 100
(3)
32
33