Environmental Research 136 (2015) 331–342

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Personal care products and steroid hormones in the Antarctic coastal environment associated with two Antarctic research stations, McMurdo Station and Scott Base Philipp Emnet a, Sally Gaw a,n, Grant Northcott b, Bryan Storey c, Lisa Graham a,1 a

Department of Chemistry, University of Canterbury, Christchurch 8140, Private Bag 4800, New Zealand Northcott Research Consultants Limited, Hamilton 3200, New Zealand c Gateway Antarctica, University of Canterbury, Christchurch 8140, Private Bag 4800, New Zealand b

art ic l e i nf o

a b s t r a c t

Article history: Received 23 July 2014 Received in revised form 8 October 2014 Accepted 10 October 2014

Pharmaceutical and personal care products (PPCPs) are a major source of micropollutants to the aquatic environment. Despite intense research on the fate and effects of PPCPs in temperate climates, there is a paucity of data on their presence in polar environments. This study reports the presence of selected PPCPs in sewage effluents from two Antarctic research stations, the adjacent coastal seawater, sea ice, and biota. Sewage effluents contained bisphenol-A, ethinylestradiol, estrone, methyl triclosan, octylphenol, triclosan, and three UV-filters. The maximum sewage effluent concentrations of 4-methyl-benzylidene camphor, benzophenone-1, estrone, ethinylestradiol, and octylphenol exceeded concentrations previously reported. Coastal seawaters contained bisphenol-A, octylphenol, triclosan, three paraben preservatives, and four UV-filters. The sea ice contained a similar range and concentration of PPCPs as the seawater. Benzophenone-3 (preferential accumulation in clams), estradiol, ethinylestradiol, methyl paraben (preferential accumulation in fish, with concentrations correlating negatively with fillet size), octylphenol, and propyl paraben were detected in biota samples. PPCPs were detected in seawater and biota at distances up to 25 km from the research stations WWTP discharges. Sewage effluent discharges and disposal of raw human waste through sea ice cracks have been identified as sources of PPCPs to Antarctic coastal environments. & 2014 Elsevier Inc. All rights reserved.

Keywords: Antarctica PPCPs Wastewater Seawater Sea ice

1. Introduction Pharmaceuticals and personal care products including soaps, lotions, toothpastes, sunscreens, fragrances, and moisturizers can contain a range of active ingredients that are collectively referred to as PPCPs (Brausch and Rand, 2011). The main inputs of PPCPs into the environment are industrial and household sewage (Daughton and Ternes, 1999; Ternes et al., 2004). PPCPs are increasingly being recognised as ubiquitous contaminants in freshwater and marine ecosystems (Klosterhaus et al., 2013; Luo et al., 2014). Antarctica is acknowledged as one of the few remaining places on earth relatively untouched by humans. The majority of scientific research stations on the Antarctic continent are located adjacent to the coast, where their industrial and domestic sewage is usually released directly into the coastal seawater (Bruni et al., n

Corresponding author. Fax: þ 64 3 364 2110. E-mail address: [email protected] (S. Gaw). 1 Current address: AsureQuality, 1C Quadrant Driver, Lower Hutt, New Zealand

http://dx.doi.org/10.1016/j.envres.2014.10.019 0013-9351/& 2014 Elsevier Inc. All rights reserved.

1997; Edwards et al., 1998). Under Annex III of the Protocol on Environmental Protection to the Antarctic Treaty (Article 5), liquid sewage needs to only be macerated before discharge into the ocean. Currently, the wastewaters from 37% of permanent research stations and 69% of summer stations lack any kind of sewage treatment (Groendahl et al., 2008). Furthermore, those Antarctic research stations with wastewater treatment plants (WWTPs) are often unable to cope with the high influx of wastewater during the summer season (Groendahl et al., 2008). Operational problems and malfunctions can occur from fluctuating water inflows, frozen pipes, or reduced microbial activity within the plant due to low temperatures (Groendahl et al., 2008). In addition, research parties stationed for long periods along the coast or on the sea ice to conduct their field work are allowed to dispose of raw human waste directly into the ocean via tidal cracks in the sea ice (Newman, 2012), a practise referred to as ‘tide-cracking’. Ships are also allowed to release food waste and sewage into the ocean, but at a distance of at least 12 nautical miles from the coast or ice shelf (Barnes and Conlan, 2007). Concerns have already been raised in the Arctic regarding the environmental risks of releasing contaminants via sewage

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discharges. These concerns have arisen due to the low biodiversity, low ambient temperatures, and consequently the more vulnerable ecosystems (Gunnarsdottir et al., 2013). To date, the majority of studies of organic contaminant pollution in the Antarctic have focused on organochlorine compounds (Risebrough, 1976), PAHs (Desideri et al., 1998), polychlorinated biphenyls (PCBs) (Risebrough, 1976; Weber and Goerke, 2003), polybrominated diphenyl ethers (PBDEs) (Dickhut et al., 2012; Hale et al., 2008), and organophosphorus flame retardants (Moeller et al., 2012). Long range transport (Desideri et al., 1998; Dickhut et al., 2012) as well as the research stations themselves (Hale et al., 2008) have been identified as the sources of these pollutants. To the best of our knowledge PPCPs and hormones have not previously been investigated as environmental pollutants in the Antarctic. PPCPs such as sunscreens and moisturizers are high use products in Antarctica due to the dry atmosphere and high UV light conditions that require frequent application of sunscreen by those working outdoors. In addition large amounts of soaps, shampoos, detergents, and disinfection products are used to maintain an adequate level of hygiene to minimise the risk of disease amongst close quartered base occupants. As PPCPs are primarily designed for external use on the human body they are subjected to few if any metabolic alterations (Ternes et al., 2004), and are washed directly off the skin in an unaltered form during showering or recreational activities (Giokas et al., 2004). Removal rates for PPCPs in wastewater treatment plants are highly variable and depend on the type of treatment and the physicochemical properties of the target compounds. Reported removal rates for PPCPs from wastewater range from 12.5% to 100% (Luo et al., 2014). As a consequence large quantities of a wide variety of PPCPs may be present in wastewater entering the WWTPs of Antarctic research stations and the treated effluents subsequently released into the Antarctic aquatic environment. Once in the environment, many PPCPs degrade relatively quickly due to photo-degradation, hydrolysis, and microbial degradation processes (Caliman and Gavrilescu, 2009). However daily use combined with the wide range of consumer products in which they are present results in their continual release into the environment, which confers them a degree of pseudo-persistence (Daughton and Ternes, 1999; Munoz et al., 2008) and promotes them as chemicals of environmental concern. In addition the prevailing polar climatic conditions, in particular the cold temperatures, extended periods of darkness, and the presence of sea ice covering coastal sea waters for a large part of the year, may reduce the degradation and therefore extend the persistence of micropollutants including PPCPs in Antarctic coastal environments. Two field studies were conducted to investigate the presence of PPCPs in Antarctica. A pilot study was conducted in October during the summer research season of 2009/2010 to determine if PPCPs were present in the sewage effluents of Scott Base (New Zealand) and McMurdo Station (USA) and the receiving coastal environment, and to collect marine biota for subsequent analysis. A larger study was conducted over November/December during the summer research season of 2012/2013 to investigate the distribution of PPCPs over a wider area of coastal water around the research stations. The specific objectives of this study were to: identify and quantify PPCPs in the effluents from the WWTPs at Scott Base and McMurdo Station on Ross Island; determine the concentration and distribution of PPCPs in the seawaters of Erebus Bay which receive these WWTP discharges and to determine if PPCPs accumulate in aquatic biota living in Erebus Bay. A preliminary assessment of the potential risk detected PPCPs may pose to Antarctica's unique marine ecosystem was carried out.

2. Materials and methods 2.1. Study area Ross Island is a volcanic island situated in the McMurdo Sound region of the Ross Sea. To the south the island is bordered by permanent ice from the Ross Ice Shelf (Fig. S1, Supporting Information). The remainder of the island is surrounded by annual sea ice which begins to break up and disperse between December and February (Falconer and Pyne, 2004). The two research stations Scott Base (New Zealand) and McMurdo Station (USA) are located on Hut Point Peninsula in the southwest of Ross Island along Erebus Bay (Fig. S1, Supporting Information). Scott Base and McMurdo Station can house up to 86 and 1200 personnel respectively over summer, with a reduced population over winter (Hale et al., 2008). The sewage treatment plants of Scott Base (aerated fixed thin-film beds) and McMurdo (extended aeration system using aerobic digestion) produce approximately 17,000 L and 416,000 L of effluent per day respectively over the summer season (October–February) (ANZ, 2011; Law et al., 2006). Scott Base began to use ozone disinfection during the 09/10 season. Further details of the sewage treatment systems, the outflow volume and temperature over one research season, the population changes of Scott Base, and the regions’ prevailing ocean currents are provided in the SI. 2.2. Sample collection Detailed sample collection protocols for each season are provided in the Supporting Information (SI). A number of precautionary measures were taken to avoid contamination of samples while working under challenging environmental field conditions. All equipment and sampling bottles were wiped with methanol before sampling to remove any contaminants from the surface. The internal surfaces of sample bottles were thoroughly rinsed with methanol and ACN. A zinc-based sunscreen was required to be worn by all field personal for protection against UV light during sampling activities. Therefore during field sampling activities all equipment was handled only while wearing disposable nitrile gloves. Samples of WWTP effluent were collected into pre-cleaned amber glass bottles, capped firmly, and transported to the laboratory at Scott Base, where they were immediately acidified to pH 2 using sulphuric acid. Seawater grab samples were collected into pre-cleaned amber 4 L Winchester bottles via bore holes in the sea ice drilled using either a Kovac or Jiffy drill. A 4 L field blank sample (Milli-Q) was included within each individual sampling trip to account for any cross-contamination of samples during their collection and subsequent handling. The bottles were stored in polystyrene padded boxes during transport back to Scott Base, where they were immediately acidified to pH 2 using sulphuric acid. 2.2.1. 2009/2010 research season Sewage effluent samples were obtained from McMurdo and Scott Base at the beginning of the summer research season, between the 23rd and 31st of October 2009. Seawater grab samples were obtained from four locations; the coast off Scott Base, Winter Quarters Bay, Cape Armitage located in between the two research stations, and a reference site at Cape Evans (Fig. S1). Clams (Laternula elliptica) were collected from Winter Quarters Bay by the McMurdo SCUBA team. Sea urchins (Sterichinus neumayeri) were collected from Cape Armitage using a remote controlled mini-submarine. Fish (Trematomus bernachii) were collected from Cape Evans by ice fishing.

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333

2

1

3

5 4

8

7 6

11

10 9

14 13

12

23 17

24 22

16

20

15 18

21

19

Fig. 1. : Location of the seawater sampling sites in Erebus Bay.

2.2.2. 2012/2013 research season The treated sewage effluent from the Scott Base WWTP was sampled monthly between August 2012 and February 2013. The samples were acidified at Scott Base and were stored on ice after

collection and during transport by air to the laboratory in Christchurch. Each monthly sample included a 1 L Milli-Q trip blank. All samples were delivered to University of Canterbury and extracted within 72 h of sampling of the WWTP. The Scott Base

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WWTP effluent was also sampled daily over the course of one week (9:30 AM, 9th – 15th Dec 2013) during the 2012/2013 field season, and extracted on-site at Scott Base. Seawater samples were obtained from 24 locations across Erebus Bay, stretching between the sea ice/ice shelf boundary in the south and as far north as Cape Evans (Fig. 1, GPS location data provided in SI). Grab seawater samples were obtained on three separate rounds between the 23rd of November and the 7th of December. During the second sampling round sea ice chips produced from the ice drilling process were also collected from five locations. 2.3. Sample preparation All water samples were filtered through a GF/C Whatman filters prior to solid phase extraction (SPE), followed by a florisil clean-up step. The seawater samples contained almost no particulates. Extraction efficiencies were checked by spiking all samples with 13C analogues of target analytes. The SPE extracts were stored at 80 °C until analysis by GC–MS. Biota samples were homogenised prior to Accelerated Solvent Extraction (ASE). Extraction efficiencies were checked by spiking all samples with 13C analogues of target analytes. The ASE extracts were cleaned by SPE, florisil, and gel permeation chromatograph (GPC) prior to GC–MS analysis. Further details on the sample preparation protocols are provided in the SI. 2.4. Sample analysis All samples were analysed using the same GC–MS instrument and methodology. Samples were derivatised (BSTFA/TMSI, 98:2) prior to analysis. Target analytes were quantified against the relative response of the internal standard using a ten-point calibration curve (0, 1, 2.5, 5, 25, 50, 75, 100, and 250 μg L  1) and internal standard quantification. The internal standard BPC was used for all sewage effluent, seawater, and sea ice samples. Three further internal standards were added for the quantification of the biota samples as the complicated matrix gave rise to matrix enhancements in certain parts of the chromatograph which the single internal standard could not correct for. Due to the high surrogate recoveries the concentrations of target analytes were not corrected against the surrogate recoveries. Method performance data, including all 13C surrogate recovery data, method detection limits for every analyte, and further details on sample derivatisation and GC–MS analysis are provided in the SI. 2.5. QA/QC All sewage, sea water, sea-ice, and associated QA/QC samples were spiked with a mixture of 13C-labelled surrogate standards immediately before SPE. The QA/QC samples included field blanks, SPE blanks and milli-Q blanks. Sewage samples were extracted in triplicate, comprising a duplicate extraction of each sample together with another which was also spiked with a mixed solution of the target analytes for use as a matrix recovery sample. Two seawater samples from each sampling round were analysed in duplicate, and a further two were spiked with the full range of target analytes for use as a matrix recovery samples to assess efficacy of the SPE procedure. The recoveries of 13C-labelled surrogate compounds together with target analytes from the matrix spike samples, were used to monitor and confirm the performance of the SPE procedure (all results presented in SI). Comparative standards were dispensed at the same time the corresponding samples were spiked before extraction. Each batch of samples subjected to SPE included a SPE cartridge

blank and spike, a Milli-Q water blank and spike, sodium sulphate blank, and solvent blank. The highest concentration of target analytes in the QA/QC samples was present in the Milli-Q water blank samples. Therefore the results reported for all WWTP effluent, sea water and sea-ice samples were corrected for analyte contributions quantified in the Milli-Q water blank corresponding to each batch of extracted samples. Occasionally the relatively low amount of target analyte spiked into matrix spike sampleswas too low in relation to the very high amount naturally present in the sample, resulting in artificially high spike recoveries. Each batch of biota samples subjected to ASE extraction batches included QA/QC samples comprising either one sample duplicate and sample matrix spike, or an Ottawa sand blank and spike. The recoveries of the 13C-labelled surrogates and target analytes were used to monitor and confirm the performance of the ASE and subsequent SPE and clean up procedures (all results are presented in SI). Comparative standards were dispensed at the same time the corresponding samples were spiked before extraction. All reported results were corrected for analyte contributions quantified in the Ottawa sand blank.

3. Results and discussion 3.1. Preliminary study, 2009/2010 research season Seven analytes were detected in the Scott Base WWTP effluent, while only three analytes were detected in the McMurdo WWTP effluent (SI, Table S6). The Scott Base WWTP effluent contained the UV-filters 4-methyl-benzylidene camphor (4-MBC), 2-hydroxy-4methoxybenzophenone (BP-3), and 2,4-dihydroxybenzophenone (BP-1), the plastic monomer 2,2-bis(4-hydroxyphenyl)propane (BPA), the steroid hormone estrone (E1), the alkylphenol 4-t-octylphenol (OP), and the antimicrobial triclosan (tric). The McMurdo WWTP effluent only contained BP-3, BP-1, and BPA. The concentration of BP-3 was lower in the Scott Base WWTP effluent than the McMurdo WWTP effluent (89.7 vs. 131 ng L  1 respectively), and that of BP-1 was higher (171 vs. 7.3 ng L  1 respectively). The concentration of BPA was similar in both effluents (31.9 vs. 28.0 ng L  1 respectively). Scott Base and McMurdo utilise different treatment processes to treat their wastewater (aerated thinfilm beds vs. extended aeration), which was a likely cause of the observed differences. There may also be differences in the types of pharmaceuticals and personal care products and frequency of their use between the populations of the two research stations. All four UV filters, BPA, p-hydroxybenzoic acid methyl ester (mParaben), p-hydroxybenzoic 155 acid butyl ester (bParaben), p-hydroxybenzoic acid propyl ester (pParaben), OP, and triclosan were detected in the seawater sampled nearby the research stations, at concentrations comparable to those reported for international coastal waters adjacent to significantly greater human populations (Table 1). This raises interesting questions with regards to why similar concentrations of these PPCPs occur in such widely different environments (temperate vs. polar) with vastly different population sizes. Unexpectedly, a total of eight analytes, BP-3, BPA, methyl triclosan, mParaben, pParaben, OMC, OP, and triclosan, were detected in seawater at the Cape Evans reference site, located 25 km north and up-current from the primary sampling area. The concentrations of the detected analytes at Cape Evans were generally lower than those measured in seawater closer to the research stations. The concentrations of target analytes measured in seawater from the individual sampling sites are provided in the SI (Table S7). Cape Evans was selected as a reference site due to its distance from the research stations. The detection of target analytes in seawater at this location was unexpected, and indicates that the

P. Emnet et al. / Environmental Research 136 (2015) 331–342

335

Table 1 Concentration range of target analytes in Antarctic WWTP effluents. Analyte

Season

Detection Frequency

Range (ng L  1)

Literature range (ng L  1)

Refs.

4-MBC

09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13

1/2 13/13 2/2 13/13 2/2 13/13 2/2 13/13 0/2 1/13 NA 13/13 1/2 11/13 0/2 0/13 0/2 7/13 0/2 0/13 0/2 0/13 0/2 2/13 0/2 13/13 0/2 0/13 0/2 0/13 1/2 13/13 0/2 0/13 1/2 13/13

173–217 321–11,700 7.3–171 24.3–6830 70–131 16.7–195 22.9–31.9 4.7–986 ND 9.7–11 NAn 168–2,700 40.9–45.7 3.1–332 ND ND ND 11.5–77.8 ND ND ND ND ND 22.7–36.4 ND 19.3–42.5 ND ND ND ND 101–118 7.5–7050 ND ND 226–248 75.2–807

42–2300

Balmer et al. (2005), Gomez et al. (2009)

o2–41

Kasprzyk-Hordern et al. (2009), Negreira et al. (2009)

3–2196

Kasprzyk-Hordern et al. (2009), Rodil et al. (2009)

6–3642

Huang et al. (2012)

o0.2–83

Blanco et al. (2009), Jonkers et al. (2010)

10–200,000

Singley et al. (1974)

o0.1–147

Liu et al. (2009), Ying et al. (2002)

0.5–18

Nakada et al. (2006), Sarmah et al. (2006)

o0.3–7.5

Combalbert and Hernandez-Raquet (2010)

0.3–275

Liu et al. (2009), Nakada et al. (2006)

o0.3–1600

Gonazles-Marino et al. (2011), Jonkers et al. (2009)

2.1–423

Benijts et al. (2004), Jonkers et al. (2009)

o2–51

Kantiani et al. (2008), Lindstrom et al. (2002)

o29–3210

Jonkers et al. (2009), Lee et al. (2005)

o10–177

Balmer et al. (2005), Gomez et al. (2009)

3.7–3949

Jonkers et al. (2010), Kasprzyk-Hordern et al. (2009)

o0.5–95

Jonkers et al. (2010), Kasprzyk-Hordern et al. (2009)

10–5370

Bedoux et al. (2012)

BP-1 BP-3 BPA bParaben Cstanol E1 E2 EE2 E3 eParaben mParaben mTric NP OMC OP pParaben Tric

ND¼ not detected, NAn ¼not analysed in 2009/10

sources and distribution mechanisms of sewage derived pollutants in the marine environment are not yet fully understood. The widespread occurrence of PPCPs in the analysed seawater samples suggests a much larger area of Erebus Bay is affected by human sewage waste than was previously thought. To gain a better understanding of the distribution and temporal variability of PPCPs in the Scott Base WWTP and Erebus Bay seawaters a more comprehensive field study was carried out over the 2012/2013 summer research season. 3.2. Comprehensive study, 2012/2013 research season 3.2.1. Sewage effluent The nine most commonly detected target analytes in the Scott Base sewage effluents during the 2012/2013 season were 4-MBC, BP-1, BP-3, BPA, Cstanol, E1, methyl triclosan, OP, and triclosan. These analytes were detected in all samples, except for E1, which was not detected in the January and February effluent samples. With the exception of the faecal steroids which were not analysed in WWTP samples collected in 2009/2010, and methyl triclosan, these same analytes were also detected in the 2009/2010 Scott Base WWTP effluent samples. In addition to these nine analytes, the synthetic estrogenic steroid ethinylestradiol (EE2) was detected in all samples of the seven-day monitoring study in December, but was absent in all other samples of Scott Base WWTP effluent. The results from the analysis of monthly Scott Base WWTP sewage samples collected over a period of six months, and the daily sewage samples collected over seven days are discussed in detail in separate sections below. The range, frequency, and

concentration of detected analytes, and a comparison to reported international values are presented in Table 1. The analyte concentrations for the individual effluent samples are provided in the SI (Tables S10 and S11). 3.2.2. Six-month study The maximum concentrations of 4-MBC, methyl triclosan, and OP (2130, 40.6 and 4070 ng L  1 respectively) in Scott Base WWTP effluent were close to the maximum concentrations observed internationally in WWTP effluents (2300, 51, and 3949 ng L  1 respectively (Table 2). The maximum concentration of BP-1 (461 ng L  1) was an order of magnitude higher than previously reported for sewage effluent (41 ng L  1). The concentrations of the other detected target analytes fall into the middle range of previously reported international data (Table 2). The average concentrations of the nine most commonly detected target analytes over the six-month sampling period are presented in Fig. 1. Concentrations of the detected analytes did not correlate with the number of staff on base present at the time of sampling, or with the operating temperature of the WWTP (Pearson's Product–Momentum Correlation test). The concentrations of most of the analytes varied from month to month. The largest concentration fluctuations in the Scott Base WWTP effluents were observed for BP-1, OP, and triclosan. The concentrations of E1 and methyl triclosan remained relatively constant. The only target analytes demonstrating potential trends in effluent concentration were OP and 4-MBC. Concentrations of the UV filter 4-MBC increased steadily throughout the research season, increasing from 321 ng L  1 in August to 2130 ng L  1 in

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P. Emnet et al. / Environmental Research 136 (2015) 331–342

Table 2 Concentration range of target analytes in Antarctic seawater. Analyte

Season

Detection frequency

Range (ng L  1)

Literature range (ng L  1)

Ref.

4-MBC

09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13 09/10 12/13

1/10 38/48 7/10 0/48 5/10 47/48 10/10 22/48 3/10 4/48 NAn 0/48 0/10 2/48 0/10 0/48 0/10 2/48 0/10 1/48 0/10 0/48 5/10 39/48 1/10 0/48 0/10 0/48 10/10 44/48 10/10 2/48 3/10 1/48 9/10 0/48

45.1 o 3.2–5.8 o 0.8–10.3 ND 12.0–88.4 o 2.6–3.7 2.2–29.5 o 1.3–7.7 o 0.5–2.3 o 0.5–0.7 NAn ND ND o 7.0 ND ND ND o 1.4 ND o 2.0 ND ND 1.9–33.3 o 0.8–37.4 o 0.2 ND ND ND o 1.9–32.3 o 1.9–4.3 0.3–1.8 0.4–0.9 o 0.8–3.0 o 0.8 o 0.5–1.7 ND

13.1–798.7

Giokas et al. (2005), Langford and Thomas (2008)

280

Tarazona et al. (2010)

1.8–3,300

Giokas et al. (2004), Tarazona et al. (2010)

0.11–2,470

Basheer et al. (2004), Beck et al. (2005)

o 0.2–7.1

Jonkers et al. (2010)

o 10–47,500

LeBlanc et al. (1992)

0.08–85

Beck et al. (2005), Pojana et al. (2004)

o 1.0–175

Pojana et al. (2007)

0.14–75

Pojana et al. (2004), Robinson et al. (2009)

ND o 0.3–15

Beck et al. (2005), Pojana et al. (2007), Jonkers et al. (2009) Jonkers et al. (2010)

2.1–62

Jonkers et al. (2010)

BP-1 BP-3 BPA bParaben Cstanol E1 E2 EE2 E3 eParaben mParaben mTric NP OMC OP pParaben Tric

NA o 0.5–755

Pojana et al. (2007), Santiago and Kwan (2007)

7.4–389.9

Giokas et al. (2005), Langford and Thomas (2008)

o 0.04–800

Basheer et al. (2004), Beck et al. (2005)

o 0.5–7.9

Jonkers et al. (2010)

0.008–99.3

Bedoux et al. (2012), Wu et al. (2007)

ND¼ not detected, NA ¼ literature data not available, NAn ¼ not analysed in 2009/10

January. Conversely, the effluent concentration of OP was highest at 4070 ng L  1 in the August sample obtained before the start of the research season, before dropping to 496 ng L  1 in October and remaining relatively low throughout the remainder of the season. 3.2.3. Seven-day study The average concentrations of the ten commonly detected PPCPs measured in daily Scott Base WWTP effluent samples over a period of one week are presented in Fig. 2. The concentrations of 4-MBC, BP-1, BPA, coprostanol, OP, and triclosan increased over the 11 days following the collection of the December monthly sample, reaching maximum concentrations of 11,700, 6830, 986, 2700, 7050, and 807 ng L  1 respectively. These concentrations exceeded those measured in any of the WWTP samples collected during the six month study. The maximum concentrations of 4-MBC, BP-1, E1, EE2, and OP were higher than those previously reported internationally in WWTP effluents (Table 1). PPCP concentrations did not correlate with the Scott Base population over the seven-day sampling period (Pearson's Product– Momentum Correlation test). WWTP discharge volumes also did not correlate over the seven-day sampling period, but strongly correlate with population over longer periods, such as the summer research season (p 40.0001). The mass flux of PPCPs in the sewage effluent is therefore likely to also correlate with population size. The maximum concentrations of 4-MBC and OP measured in the daily sampled WWTP effluents exceeded those measured during the six-month study. This demonstrates that field studies conducted over long periods of time with low sampling frequencies can miss short-term fluctuations in the concentrations of

target analytes. The detection of EE2 in these effluent samples also demonstrates that low frequency sampling carries the risk of not capturing the release of PPCPs in WWTP effluents and consequently of underestimating their potential risk to receiving environments. The temperature of the Scott Base WWTP remained constant at  24 °C over the week in which daily sampling was completed. The fluctuation in the concentration of measured PPCPs can therefore not be ascribed to changes in the operating temperature of the WWTP. The observed analyte concentration fluctuations are likely a reflection of the constantly and rapidly fluctuating volume and composition of influent entering the WWTP. The influent composition can be predominantly comprised of domestic grey water one day, and be heavily influenced by waste streams from the mechanical workshops the next day. When science field parties return to base their liquid field waste is disposed into the WWTP, adding further pulses of waste. 3.2.4. Ocean water This study is the first to report on the presence of a range of PPCPs in the WWTP sewage effluent of two Antarctic research stations and the receiving coastal environment. PPCPs were frequently detected in the seawaters of Erebus Bay, including the reference site at a distance of 25 km up-current from the identified point sources. The target analytes detected in the Erebus Bay seawaters during the 2012/2013 research season were similar to those detected during the 2009/2010 research season, and lie in the lower range of concentrations reported internationally in coastal and ocean waters (Table 2). Eleven analytes were detected in the Erebus Bay seawater samples, namely 4-MBC, BPA, BP-3,

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337

Fig. 2. : Average analyte concentrations (ng L  1) of the most frequently detected PPCPs in Scott Base sewage effluents during the six month (a) and seven day (b) sampling periods. Analytes have been grouped according to measured concentration ranges. * indicates when the ozonation plant at Scott Base was in operation. Lines on (b) indicate maximum concentrations measured during the six month study (except for EE2, which was not previously detected).

bParaben, E1, EE2, E3, mParaben, OMC, OP, and pParaben (Table 2). The most commonly detected analytes were 4-MBC, BP-3, BPA, mParaben, and OMC (Fig. 3). The less commonly detected analytes EE2, E3, and OP were only detected in seawater at sites close to the research stations. E1, EE2, E3, and pParaben were only detected below the MQL. The detected analyte concentrations from each sampling location and sampling round are provided in the SI (Table S14). Lower concentrations of PPCPs were measured in seawater sampled in 2012/2013 compared to 2009/2010 (Table 2). There are two possible explanations for this observation. Firstly, the sea ice in Erebus Bay had broken out in early 2012 for the first time in over 13 years (Szorc et al., 2011). The long-term presence of the sea ice may have affected the behaviour and distribution of PPCPs in Erebus Bay. The presence of sea ice has also been suggested to alter the degradation behaviour of PPCPs in sea ice dominated environments, possibly prolonging their presence in the environment (Emnet et al., 2013). Secondly, the field work completed over the 2012/2013 season was carried out approximately 6 weeks later in the season than in 2009/2010. The seawater concentrations of the most commonly detected analytes 4-MBC, BP-3, BPA, mParaben, and OMC remained relatively constant throughout Erebus Bay. This suggests the seawater is well mixed, but that limited dilution of the sewage effluents occurs as it becomes distributed across Erebus Bay towards Cape Evans. Consequently the concentration of detected analytes did not differ between the three sampling rounds completed in 2012/2013.

As observed during the 2009/2010 field study, the analytes mParaben and OMC were detected in the majority of seawater samples collected in the 2012/2013 season, but were not detected in the WWTP effluents discharged from Scott Base. The QA/QC blanks indicated only occasional minor background contamination by BP-3, mParaben, OP, and OMC, which was conservatively subtracted from the environmental data. Poor extraction efficiencies of mParaben and OMC from the sewage effluents can be excluded as a possible explanation of this observation as the recovery of surrogate and target analytes spikes was within acceptable ranges. The disposal of raw human waste and grey water from field research parties via tidal cracks in the sea ice may be a source of mParaben and OMC. The practise of waste disposal by ‘tidecracking’ is permitted under some circumstances (Newman, 2012). Field parties stationed for long periods (up to several months) on or near the sea ice are permitted to dispose of their wastes by tide cracking because of the impracticality of returning large volumes of waste to the bases for treatment (Newman, 2012). This practice is only subject to environmental impact assessments, and therefore regulated and monitored, and avoided whenever possible. However both the US and New Zealand Antarctic programs have reported the disposal of several thousand litres of human waste and grey water in a single season at a number of sites throughout the McMurdo Sound and Erebus Bay by tide-cracking (Newman, 2012; Pettway, 2012). This includes wastewater, including toilet waste, generated by the airport and ice runway operated by the US

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Fig. 3. Distribution of (a) bis-phenol. A and (b) methyl paraben in seawater for the 2012/13 season. The horizontal lines represent the method quantification limits.

Antarctic Programme. Since this source of human waste is untreated it may contain PPCPs that would have been removed during wastewater treatment. Tide-cracked waste may therefore be a source of the parabens and OMC that were detected in the coastal seawater samples but not in the discharged WWTP effluents. 3.2.5. Sea Ice The potential for sea ice to act as an environmental sink and potential source of PPCPs within Erebus Bay was investigated by determining the concentration of PPCPs present in the sea ice. The concentration range of target analytes detected in the sea ice is provided in Table 3. As no data exist on the occurrence of PPCPs in sea ice the detected concentrations were compared to previous studies reporting their concentrations in seawater. The concentrations detected in sea ice from each of the five sampling locations are provided in the SI (Table S15). A total of seven target analytes were detected in the sea ice, namely 4-MBC, BP-3, BPA, E1, EE2, OMC, and OP (Table 3). With the exception of E1, EE2, and OP these were also the most commonly detected target analytes in the seawater within Erebus Bay. However mParaben was an exception as it was not detected in the

Table 3 Concentration range of target analytes in Antarctic sea ice. Analyte

Season

Detection frequency

Range (ng L  1)

4-MBC BP-1 BP-3 BPA bParaben Cstanol E1 E2 EE2 E3 eParaben mParaben mTric NP OMC OP pParaben Tric

12/13 12/13 12/13 12/13 12/13 12/13 12/13 12/13 12/13 12/13 12/13 12/13 12/13 12/13 12/13 12/13 12/13 12/13

5/5 0/5 5/5 2/5 0/5 0/5 2.5 0/5 1/5 0/5 0/5 0/5 0/5 0/5 5/5 2/5 0/5 0/5

o 3.2–4.3 ND o 2.6–3.8 o 1.3 ND ND o 7.0 ND o 1.4 ND ND ND ND ND o 1.9–4.8 0.5–0.9 ND ND

ND¼ not detected

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sea ice, despite being frequently detected in the underlying seawater. The concentration range of the detected target analytes in the sea ice samples is comparable to that detected in the corresponding seawater samples. The PPCPs present in the sea ice will have likely been present in seawater during the previous research season and become entrapped in the sea ice as it re-formed during autumn and winter (Falconer and Pyne, 2004; Perovich, 1993). During the following season these PPCPs would be released back into the seawater as the ice melts during summer. The sea ice may also transport these PPCPs to previously un-impacted areas when it breaks out and is transported away by the ocean currents. Large otherwise isolated areas of the aquatic environment may therefore be exposed to seasonal pulses of sewage-derived PPCPs. 3.2.6. Biota Matrix interferences prevented the detection and/or quantification of a number of spiked surrogates and target analytes from the biota samples. The surrogates 13C6-bParaben, 13C6-E2, and 13 C6-mParaben, and target analytes BP-1, BP-3, bParaben, Cstanol, E1, E2, EE2, E3, eParaben, mParaben, OP, and pParaben could be adequately quantified. A total of six PPCPs (BP-3, E2, EE2, mParaben, OP, and pParaben) and the faecal steroid coprostanol could be detected in biota from Erebus Bay (Table 4). The detected analyte concentrations from each biota sample are provided in the SI (Table S16). The analytes concentrations are reported by dry weight (d.w.), wet weight (w.w.), and lipid weight (l.w.). Lipid contents of the clams and urchin composite were too low to be determined accurately by gravimetric methods. The lipid content of the fish tissues was also low as the fish were caught in the beginning of spring when they are just beginning to feed after their winter starvation period (Kawaguchi et al., 1989). The reported lipid based concentrations are therefore elevated and need to be considered carefully when making comparisons to previous studies.

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The highest concentrations of BP-3 were found in clam tissues, while mParaben concentrations were highest in fish muscle tissues. The clams obtained from Winter Quarters Bay off the coast of McMurdo Station contained up to 112, 10.9, 23.1, 5.8, and 5.3 ng g  1 d.w. of BP-3, E2, EE2, mParaben, and pParaben respectively. The sea urchin composite sample from Cape Armitage (located between McMurdo Station and Scott Base) contained 8.6 and 5.7 ng g  1 d.w. of BP-3 and mParaben respectively. The human faecal sterol indicator coprostanol was also be detected in five out of the seven analysed clams, and the urchin composite. Interestingly, the fish, obtained 25 km distance from the two research stations at Cape Evans, contained up to 14.1 19.2, and 5.0 ng g  1 d. w. of BP-3, mParaben, and OP respectively. T. bernachii do not migrate beyond 500 m of their point of release after tagging (Kawaguchi et al., 1989). The limited migration range for the fish either indicates an unaccounted source of PPCPs, or that PPCPs from the research stations continuously circulate to the seawaters at Cape Evans, allowing for sufficient exposure time for the PPCPs to bioaccumulate. Tide-cracked waste would only provide brief pulses of PPCPs. The fish liver contained 41.0 and 2.4 ng g  1 d.w. BP-3 and mParaben respectively, indicating the liver preferentially accumulates BP-3, but not mParaben and OP, compared to the fish muscle tissue. The wide geographical distribution of organic contaminants within Erebus Bay has been previously observed (Hale et al., 2008). Detectable concentrations of polybrominated diphenyl ether flame retardants (PBDEs) have been measured in the clam L. elliptica and the fish T. bernachii (the same species used here) as far as 25 km away from the research bases (Hale et al., 2008). PBDEs were also detected in marine sediments from and nearby to Winter Quarters Bay (Hale et al., 2008). Our results provide the first example of paraben preservatives bioaccumulating in a field study. The evidence for bioaccumulation of mParaben and pParaben is unexpected due to their low log KOW values of 1.66 and 2.71 respectively (Golden et al., 2005). However, other mechanisms such as partitioning into non-water and non-lipid

Table 4 Concentration range (and detection frequency) and comparison to literature values of detected target analytes in Antarctic clams (n ¼7), urchin composite (n¼10), fish (n¼ 7), and fish liver. Concentrations are reported in ng g  1 wet weight (w.w.), ng g  1 dry weight (d.w.), and ng g  1 lipid weight (l.w.) to allow for comparisons with the literature. Analyte

Clams

Urchin composite

Fish

Fish liver

Literature range

Refs.

BP-3

d.w. w.w. l.w.

9.2–112 (7) 1.4–23.1 n

8.6 0.9 n

o 6.6–14.1 o 1.3–3.0 265–1450

41 9.6 1690

11.2–24.3 22–298

Gago-Ferrero et al. (2013)

Cstanol

d.w. w.w. l.w.

76.2–230 (5) 9.4–35.9 n

1260 133 n





NA

NA

E2

d.w. w.w. l.w.

5.1–10.9 (3) 0.8–2.0 n







NA

NA

EE2

d.w. w.w. l.w.

8.1–23.1 (4) 1.5–4.3 n







o 3–38

Pojana et al. (2007)

mParaben

d.w. w.w l.w.

o2.1–5.8 (7) o0.4–1.0 n

5.7 0.6 n

5.1–26.9 1.0–6.1 202–1,990

2.4 0.3 100

860–2300

Barse et al. (2010)

OP

d.w. w.w. l.w.





1.6–5.0 0.3–1.1 63.9–464

– 2.7–44.9 3190–4920

Ferrara et al. (2001), Basheer et al. (2004) Bennett and Metcalfe (2000)

d.w. w.w. l.w.

2.1–5.3 (4) 0.4–1.9 n



– 870–6700

Bjerregaard et al. (2003)

pParaben



Nagtegaal et al. (1997), Balmer et al. (2005), Fent et al. (2010)

a biota reference values include fish and shellfish. NA ¼no literature data available nfat content of the biota tissue was too low to be accurately determined gravimetrically.

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cellular components of the organism may play a role in bioaccumulation (Klosterhaus et al., 2013). These mechanisms include involve transport processes rather than diffusion processes, and sorption to proteins (Klosterhaus et al., 2013). This bioaccumulation may also be highly species and compound specific, and may not necessarily occur in other Antarctica biota. For example, the bioaccumulation of various classes of organochlorine chemicals has been previously observed to be species specific (Pastor et al., 1996). These results suggest that physiochemical properties such as log KOW cannot be solely relied upon to predict whether a contaminant may be of environmental concern. This suggests other contaminants which are not currently considered to present an environmental risk on the basis of their physicochemical properties may ultimately prove to be of environmental concern. There are limited data reporting the presence of the PPCPs detected in Antarctic marine animals in aquatic biota around the world (Table 2). mParaben displayed dose dependant accumulation in the gill, liver, muscle, brain, and testis tissue of the freshwater fish Cyprinus carpio (Barse et al., 2010). pParaben accumulated in the liver and muscle tissue of the freshwater fish Oncorhynchus mykiss (Bjerregaard et al., 2003). The concentration of OP detected in the Cape Evans fish (0.3–1.1 ng g  1 w.w.) lies in the lower ranges of concentrations reported in aquatic biota, including fish (2.7–44.9 ng g  1 w.w.) (Basheer et al., 2004;,Ferrara et al., 2001). Bioaccumulation of OP has been reported in fish (Ferreira-Leach and Hill, 2001; Ying, 2006) and shellfish (Bennett and Metcalfe, 2000). While OP can be metabolised and excreted via the liver/bile route by fish, OP has been shown to accumulate in fish brain, muscle, skin, bone, gill, and eye tissue (Ferreira-Leach and Hill, 2001). The dry weight based concentration range of BP-3 (4.7–112 ng g  1 d.w.) measured in the Antarctic fish is up to 5-fold higher than that previously reported in fish (11.2–24.3 ng g  1 d. w.) (Gago-Ferrero et al., 2013). The lipid based BP-3 concentration ranges in the Cape Evans fish (265–1450 ng g  1 l.w.) are equally elevated compared to previous studies in fish from around the world ( o50–298 ng g  1 l.w.) (Balmer et al., 2005; Fent et al., 2010; Gago-Ferrero et al., 2013; Nagtegaal et al., 1997; Zenker et al., 2008). This is likely in part due to Antarctic fish lipid contents being low following the winter starvation period. It is likely the lipid-based concentrations of the detected PPCPs in fish would be reduced at the end of summer when fish have been feeding and their reserves of lipid have increased. The concentrations of EE2 measured in Antarctic clams (8.1–23.1 ng g  1 d.w.) lie in the same range as reported internationally in mussels (o3–38 ng g  1 d.w.) (Pojana et al., 2007).

4. Environmental implications These results for PPCPs in Antarctic research base wastewater discharges and the surrounding coastal environment in Erebus Bay are likely to be replicated at other Antarctic research stations. A number of PPCPs exhibit endocrine disrupting properties (Daughton and Ternes, 1999). For example, UV filters are known to exhibit estrogenic activity (Brausch and Rand, 2011; Fent et al., 2008), but have also been demonstrated to confer antiestrogenic, androgenic, and antiandrogenic responses (Brausch and Rand, 2011). The estrogenic activity of parabens have been shown to increase with increasing length of the alkyl chain (Golden et al., 2005). OP can cause endocrine disrupting effects in marine and freshwater species (Ying, 2006), triclosan is toxic to algae, microorganisms, amphibians, and fish larvae (Bedoux et al., 2012), and BPA can cause toxic and endocrine disrupting effects in aquatic organisms (Kang et al., 2007). The presence of steroid hormone residues in WWTP effluent released into Antarctic coastal waters

is of concern due to their strong biological potency they exhibit at low ng L  1 concentrations (Liu et al., 2009; Mes et al., 2005). Similarly, triclosan and BPA can elicit biological effects at very low environmental concentrations (Bedoux et al., 2012; Kang et al., 2007). None of the commonly target analytes were detected in the seawaters at concentrations above 88 ng L  1. These maximum aqueous concentrations of detected PPCPs were orders of magnitude lower than those reported to induce biological effects, including growth endpoints, endocrine disruption and acute toxicity (Baptista et al., 2009; Brausch and Rand, 2011; Gago-Ferrero et al., 2012; Jobling et al., 1996; Kang et al., 2007; Routledge et al., 1998; Schlumpf et al., 2001; Terasaki et al., 2009; Watts et al., 2001). The more biologically active steroid hormones E1, EE2, and E3 were detected only infrequently at concentrations close to the quantification limits. Nonetheless, the potential risk these chemicals pose to Antarctic's unique marine ecosystem remains unknown. The unique climatic conditions that prevail in Antarctica, principally the cold ocean temperatures ( 0.35 to  1.9 °C), and extended periods of darkness and presence of sea ice covering coastal sea waters for a large part of the year, may reduce degradation processes such as microbial and photo-degradation, and therefore enhance the persistence of PPCPs within the unique Antarctic coastal marine environment (Emnet et al., 2013). A larger than expected area around the research stations, and therefore potentially a large number of animals within Erebus Bay, are impacted by sewage discharges. Antarctic biota generally have very slow metabolisms and are slow growing (Peck, 2002). This may reduce the excretion rates of potentially harmful chemicals, therefore leading to longer in vivo exposure periods. Critical periods of biological development may therefore also be longer than in other aquatic organisms in different regions of the world, with the consequence that endocrine disruption may have a particularly severe detrimental effect. Meroplankton, the embryonic life-stage of many marine organisms, has been found in Erebus Bay throughout the year, including winter (Sewell and Jury, 2011), when degradation processes of PPCPs which may occur in summer, are reduced. Antarctic biota may therefore be particularly sensitive to the effects of PPCPs. A large proportion of Antarctic biota live in, or in close proximity to marine sediments. These bottom dwelling organisms throughout Erebus Bay are therefore likely exposed to PPCPs via the sediments. The biologically active PPCPs mParaben, pParaben, BP-3, E2 and EE2 were detected in the sediment dwelling clams. Animals feeding on these clams will therefore be exposed to estrogenic chemicals in the lower ng g  1 range (wet weight basis). Other sediment dwelling organisms such as polychaete worms which are the main food source for some Antarctic fish, may also accumulate these PPCPs. Predatory species such as seals feeding on contaminated fish may also become exposed to micropollutants.

5. Implications for future research Our results have shown that PPCPs are present in the Antarctic environment at concentrations similar to those reported elsewhere in the world. The fate of these PPCPs in the Antarctic marine environment remains largely unknown. More extensive sampling of the WWTP discharges of Scott Base and McMurdo Station during the year is required to gain an improved understanding of the mass load of PPCPs being released into the adjacent Antarctic coastal waters. Analysis of both the dissolved and particulate phases of discharged effluents is required as a significant portion of some PPCPs may adsorb to the particulate phase and would not have been measured in this study. Inclusion of the particulate

P. Emnet et al. / Environmental Research 136 (2015) 331–342

phase would allow for an estimate of the PPCPs mass loadings of sewage effluents, which could be compared to the masses measured in the seawaters of Erebus Bay to determine if the WWTP discharges are the main source of PPCPs in Erebus Bay. The contributions of tide-cracking and the PPCP storage capacity of sea ice could then be estimated. Assessment of the accumulation of PPCPs in marine sediments is also required along with their potential to transfer to sediment-dwelling organisms. The bioaccumulation mechanism (ingestion, exposure to sediments, bioconcentration from seawater) should also be investigated. Further research is needed on seasonal changes in concentration with respect to changes in sunshine, effluent volumes, and sea ice, which may influence the flow of ocean currents. This research should also be extended to other PPCPs and pharmaceuticals which can logically be expected to be present in the WWTP effluents and seawater. It should be determined whether some PPCPs could be used as WWTP specific indicator chemicals to use as tracers to study the flow, dispersion, and dilution of the effluent plumes to help verify the source of the PPCPs detected at Cape Evans. These data could also be used to identify risk zones for future studies to assess the potential accumulation in sediments and impacts on biota.

Notes The authors declare no competing financial interest.

Acknowledgements We thank Antarctica New Zealand for providing funding and logistical support for the fieldwork component of this study, made possible through the Christchurch City Council Scholarship. The authors also thank the Keith Laugesen PhD Scholarship for funding support for Philipp Emnet. We thank Professor William Davison, and Professor Ian Shaw for their guidance with this project. The authors thank Dr Miles Lamare (Otago University) and his research team for incorporating this fieldwork into their own field activities during the 2009/10 season. Matt Walters is thanked for assistance with preparing figures.

Appendix A. Suplementary Information Supplementary data associated with this article can be found in the online version at http://dx.doi.org/10.1016/j.envres.2014.10. 019.

References ANZ ANZ. Waste Management and Remediation Christchurch; 26 October 2011. Balmer, M.E., Buser, H.-R., Mller, M.D., Poiger, T., 2005. Occurrence of some organic UV filters in wastewater, in surface waters, and in fish from Swiss lakes. Environ. Sci. Technol. 39, 953–962. Baptista, M.S., Stoichev, T., Basto, M.C.P., Vasconcelos, V.M., 2009. Vasconcelos MTSD. Fate and effects of octylphenol in a Microcystis aeruginosa culture medium. Aquat. Toxicol. 92, 59–64. Barnes, D.K.A., Conlan, K.E., 2007. Disturbance, colonization and development of Antarctic benthic communities. Philos. Trans. R. Soc. B 362, 11–38. Barse, A.V., Chakrabarti, T., Ghosh, T.K., Pal, A.K., Kumar, N., Raman, R.P., Jadhao, S.B., 2010. Vitellogenin induction and histo-metabolic changes following exposure of Cyprinus carpio to methyl paraben. Asian-Australas. Assoc. Anim. Prod. Soc. 23, 1557–1565. Basheer, C., Lee, H.K., Tan, K.S., 2004. Endocrine disrupting alkylphenols and bisphenol A in coastal waters and supermarket seafood from Singapore. Mar. Pollut. Bull. 48, 1145–1167.

341

Beck, I.-C., Bruhn, R., Gandrass, J., Ruck, W., 2005. Liquid chromatography–tandem mass spectrometry analysis of estrogenic compounds in coastal surface water of the Baltic Sea. J. Chromatogr. A 1090, 98–106. Bedoux, G., Roig, B., Thomas, O., Dupont, V., Bot, B.L., 2012. Occurrence and toxicity of antimicrobial triclosan and by-products in the environment. Environ. Sci. Pollut. Res. 19, 1044–1065. http://dx.doi.org/10.1007/s11356-011-0632-z. Benijts, T., Lambert, W., Leenheer, A.D., 2004. Analysis of multiple endocrine disruptors in environmental waters via wide-spectrum solid phase extraction and dual-polarity ionization LC-Ion Trap-MS/MS. Anal. Chem. 76, 704–711. Bennett, E.R., Metcalfe, C.D., 2000. Distribution of degradation products of alkylphenol ethoxylates near sewage treatment plants in the lower Great Lakes, North America. Environ. Toxicol. Chem. 19, 784–792. Bjerregaard, P., Andersen, D.N., Pedersen, K.L., Pedersen, S.N., Korsgaard, B., 2003. Estrogenic effect of propylparaben (propylhydroxybenzoate) in rainbow trout Oncorhynchus mykiss after exposure via food and water. Comp. Biochem. Physiol. Part C 136, 309–317. Blanco, E., MdC, Casais, MdC, Mejuto, Cela, R., 2009. Combination of off-line solidphase extraction and on-column sample stacking for sensitive determination of parabens and p-hydroxybenzoic acid in waters by non-aqueous capillary electrophoresis. Anal. Chim. Acta 647, 104–111. Brausch, J.M., Rand, G.M., 2011. A review of personal care products in the aquatic environment. Environ. Conc. Tox. Chemosphere 82, 1518–1532. Bruni, V., Maugeri, T.L., Monticelli, L., 1997. Faecal pollution indicators in the Terra Nova Bay (Ross Sea, Antarctica). Mar. Pollut. Bull. 34, 908–912. Caliman, F.A., Gavrilescu, M., 2009;. Pharmaceuticals, personal care products and endocrine disrupting agents in the environment – a review. Clean 37, 277–303. Combalbert, S., Hernandez-Raquet, G., 2010. Occurrence, fate, and biodegradation of estrogens in sewage and manure. Appl. Microbiol. Biotechnol. 86, 1671–1692. Daughton, C.G., Ternes, T.A., 1999. Pharmaceuticals and personal care products in the environment: agents of subtle change? Environ. Health Perspect. 107, 907–938. Desideri, P.G., Lepri, L., Udisti, R., Checchini, L., Bubba, M.D., Cini, R., Stortini, A.M., 1998. Analysis of organic compounds in Antarctic snow and their origin. Int. J. Environ. Anal. Chem. 71, 331–351. Dickhut, R.M., Cincinelli, A., Corchan, M., Kylin, H., 2012. Aerosol-mediated transport and deposition of brominated diphenyl ethers to Antarctica. Environ. Sci. Technol. 46, 3135–3140. Edwards, D.D., McFeters, G.A., Venkatesan, M.I., 1998. Distribution of Clostridium perfringens and fecal sterols in a benthic coastal marine environment influenced by the sewage outfall from McMurdo Station, Antarctica. Appl. Environ. Microbiol. 64, 2596–2600. Emnet, P., Kookana, R.S., Shareef, A., Gaw, S., Williams, M., Crittenden, D., Northcott, G.L., 2013. The effect of irradiance and temperature on the role of photolysis in the removal of organic micropollutants under Antarctic conditions. Environ. Chem. 10, 417–423. Falconer, T.R., Pyne, A.R., 2004. ce breakout history in Southern McMurdo Sound, Antarctica (1988–2002).. ftp://ftp.geo.vuw.ac.nz/ARC/: School of Earth Sciences, Victoria University, Wellington (March). Fent, K., Kunz, P.Y., Gomez, E.U.V., 2008. Filters in the Aquatic Environment Induce Hormonal Effects and Affect Fertility and Reproduction in Fish. Chimica 62, 368–375. Fent, K., Zenker, A., Rapp, M., 2010. Widespread occurrence of estrogenic UV-filters in aquatic ecosystems in Switzerland. Environ. Pollut. 158, 1817–1824. Ferrara, F., Fabietti, F., Delise, M., Bocca, A.P., Funari, E., 2001. Alkylphenolic compounds in edible molluscs of the Adriatic Sea (Italy). Environ. Sci. Technol. 35, 3109–3112. Ferreira-Leach, A.M.R., Hill, E.M., 2001. Bioconcentration and distribution of 4-tertoctylphenol residues in tissues of the rainbow trout (Oncorhynchus mykiss). Mar. Environ. Res. 51, 75–89. Gago-Ferrero, P., Diaz-Cruz, M.S., Barcelo, D., 2012. An overview of UV-absorbing compounds (organic UV filters) in aquatic biota. Anal. Bioanal. Chem. 404, 2597–2610. Gago-Ferrero, P., Diaz-Cruz, M.S., Barcelo, D., 2013. Multi-residue method for trace level determination of UV filters in fish based on pressurized liquid extraction and liquid chromatography-quadrupole-linear ion trap-mass spectrometry. J. Chromatogr. A 1286, 93–101. Giokas, D.L., Sakkas, V.A., Albanis, T.A., 2004. Determination of residues of UV filters in natural waters by solid-phase extraction coupled to liquid chromatography– photodiode array detection and gas chromatography–mass spectrometry. J. Chromatogr. A 1026, 289–293. Giokas, D.L., Sakkas, V.A., Albanis, T.A., Lampropoulou, D.A., 2005. Determination of UV-filter residues in bathing waters by liquid chromatography UV-diode array and gas chromatography–mass spectrometry after micelle mediated extraction-solvent back extraction. J. Chromatogr. A 1077, 19–27. Golden, R., Gandy, J., Vollmer, G., 2005. A review of the endocrine activity of parabens and implications for potential risks to human health. Crit. Rev. Toxicol. 35, 435–458. Gomez, M.J., Gomez-Ramos, M.M., Aguera, A., Mezcua, M., Herrera, S., FernandezAlba, A.R., 2009. A new gas chromatography/mass spectrometry method for the simultaneous analysis of target and non-target organic contaminants in waters. J. Chromatogr. A 1216, 4071–4082. Gonazles-Marino, I., Quintana, J.B., Rodriguez, I., Cela, R., 2011. Evaluation of the occurrence and biodegradation of parabens and halogenated by-products in wastewater by accurate- mass liquid chromatography-quadrupole-time-offlight-mass spectrometry (LC–QTOF–MS). Water Res. 45, 6770–6780.

342

P. Emnet et al. / Environmental Research 136 (2015) 331–342

Groendahl, F., Sidenmark, J., Thomsen, A., 2008. Survey of waste water disposal practices at Antarctic research stations. Polar Res. 28, 298–306. Gunnarsdottir, R., Jenssen, P.D., Jensen, P.E., Villumsen, A., Kallenborn, R., 2013. A review of wastewater handling in the Arctic with special reference to pharmaceuticals and personal care products (PPCPs) and microbial pollution. Ecol. Eng. 50, 76–85. Hale, R.C., Kim, S.L., Harvey, E., Guardia, M.J.L., Mainor, T.M., Bush, E.O., Jacobs, E.M., 2008. Antarctic research bases: local sourcse of polybrominated diphenyl ether (PBDE) flame retardants. Environ. Sci. Technol. 42, 1452–1457. Huang, Y.Q., Wong, C.K.C., Zheng, J.S., Bouwman, H., Barra, R., Wahlstroem, B., Neretin, L., Wong, M.H., 2012. Bisphenol A (BPA) in China: a review of sources, environmental levels, and potential human health impacts. Environ. Int. 42, 91–99. Jobling, S., Sheahan, D., Osborne, J.A., Matthiessen, P., Sumpter, J.P., 1996. Inhibition of testicular growth in rainbow trout (Oncorhynchus mykiss) exposed to estrogenic alkylphenolic chemicals. Environ. Toxicol. Chem. 15, 194–202. Jonkers, N., Kohler, H.-P.E., Dammshauser, A., Giger, W., 2009. Mass flows of endocrine disruptors in the Glatt River during varying weather conditions. Environ. Pollut. 157, 714–723. Jonkers, N., Sousa, A., Galante-Oliveira, S., Barroso, C.M., Kohler, H.-P.E., Giger, W., 2010. Occurence and sources of selected phenolic endocrine disruptors in Ria de Aveiro, Portugal. Environ. Sci. Pollut. Res. 17, 834–843. Kang, J.-H., Aasi, D., Katayama, Y., 2007. Bisphenol A in the aquatic environment and its endocrine-disruptive effects on aquatic organisms. Crit. Rev. Toxicol. 37, 607–625. Kantiani, L., Farre, M., Asperger, D., Rubio, F., Gonzalez, S., MJLd, Alda, Petrovic, M., Shelver, W.L., Barcelo, D., 2008. Triclosan and methyl-triclosan monitoring study in the northeast of Spain using a magnetic particle enzyme immunoassay and confirmatory analysis by gas chromatography–mass spectrometry. J. Hydrol. 361, 1–9. Kasprzyk-Hordern, B., Dinsdale, R.M., Guwy, A.J., 2009. The removal of pharmaceuticals, personal care products, endocrine disruptors and illicit drugs during wastewater treatment and its impact on the quality of receiving waters. Water Res. 43, 363–380. Kawaguchi, K., Ishikawa, S., Matsuda, O., Naito, Y., 1989. Tagging experiments of nototheniid fish, Trematomus bernachii Boulenger under the coastal fast ice in Lutzow-Holm Bay, Antarctica. Proc. NIPR Symp. Polar Biol. 2, 111–116. Klosterhaus, S.L., Grace, R., Hamilton, M.C., Yee, D., 2013. Method validation and reconnaissance of pharmaceuticals, personal care products, and alkylphenols in surface waters, sediments, and mussels in an urban estuary. Environ. Int. 54, 92–99. Langford, K.H., Thomas, K.V., 2008. Inputs of chemicals from recreational activities into the Norwegian coastal zone. J. Environ. Monit. 10, 894–898. Law, K.P., Kulchawik, R.J., Zenz, D.R., Bouchard, T.B., 2006. Overview of the Design, Construction, and Operation of the Mcmurdo Wastewater Treatment Facility in Antarctica. Water Environment Foundation (WEFTEC), Chicago. LeBlanc, L.A., Latimer, J.S., Ellis, J.T., Quinn, J.G., 1992. The geochemistry of coprostanol in waters and surface sediments from Narragansett Bay. Estuar. Coast. Shelf Sci. 34, 439–458. Lee, H.-B., Peart, T.E., Svoboda, M.L., 2005. Determination of endocrine-disrupting phenols, acidic pharmaceuticals, and personal-care products in sewage by solid-phase extraction and gas chromatography–mass spectrometry. J. Chromatogr. A 1094, 122–129. Lindstrom, A., Buerge, I.J., Poiger, T., Bergqvist, P.-A., Muller, M.D., Buser, H.-R., 2002. Occurrence and environmental behavior of the bactericide triclosan and its methylated derivative in surface waters and in wastewater. Environ. Sci. Technol. 36, 2322–2329. Liu, Z.-H., Kanjo, Y., Mizutani, S., 2009. Urinary excretion rates of natural estrogens and androgens from humans, and their occurrence and fate in the environment: a review. Sci. Total Environ. 407, 4975–4985. Luo, Y., Guo, W., Ngo, H.H., Nghiem, L.D., Hai, F.I., Zhang, J., Liang, S., Wang, X.C., 2014. A review on the occurrence of micropollutants in the aquatic environment and their fate and removal during wastewater treatment. Sci. Total Environ. 473-474, 619–641. Mes, T.Z.Dd, Zeeman, G., Lettinga, G., 2005. Occurence and fate of estrone, 17bestradiol and 17a-ethynylestradiol in STPs for domestic wastewater. Rev. Environ. Sci. Biotechnol. 4, 275–311. Moeller, A., Sturm, R., Xie, Z., Cai, M., He, J., Ebinghaus, R., 2012. Organophosphorus flame retardants and plasticizers in airborne particles over the Northern Pacific and Indian Ocean toward the polar regions: evidence for global occurrence. Environ. Sci. Technol. 46, 3127–3134. Munoz, I., Gomez, M.J., Molina-Diaz, A., Huijbregts, M.A.J., Fernandez-Alba, A.R., Garcia-Calvo, E., 2008. Ranking potential impacts of priority and emerging pollutants in urban wastewater through life cycle impact assessment. Chemosphere 74, 37–44. Nagtegaal, M., Ternes, T.A., Baumann, W., Nagel, R., 1997. UV-Filtersubstanzen in Wasser und Fischen. Z. Umweltchem. Okotox. 9, 79–86.

Nakada, N., Tanishima, T., Shinohara, H., Kiri, K., Takada, H., 2006. Pharmaceutical chemicals and endocrine disrupters in municipal wastewater in Tokyo and their removal during activated sludge treatment. Water Res. 40, 3297–3303. Negreira, N., Rodriguez, I., Ramil, M., Rubi, E., Cela, R., 2009. Sensitive determination of salicylate and benzophenone type UV filters in water samples using solidphase microextraction, derivatization and gas chromatography tandem mass spectrometry. Anal. Chim. Acta 638, 36–44. Newman, J., 2012. Human waste disposal practices in Antarctica by Antarctica New Zealand. In: Emnet, P. (Ed.), Personal Communcation. Christchurch, Antarctica New Zealand. Pastor, D., Boix, J., Fernandez, V., Albaiges, J., 1996. Bioaccumulation of organochlorinated contaminants in three estuarine fish species (Mullus barbatus, Mugil cephalus, and Dicentrarcus labrax). Mar. Pollut. Bull. 32, 257–262. Peck, L.S., 2002. Ecophysiology of Antarctic marine ectotherms: limits to life. Polar Biol. 25, 31–40. Perovich, D.K., 1993. A theoretical model of ultraviolet light transmission through Antarctic sea ice. J. Geophys. Res. 98, 579–587. Pettway, K., 2012. Human waste disposal practices in Antarcitca by the US Antarctic Program. In: Emnet, P. (Ed.), Personal Communication. Antarctic Program, US. Pojana, G., Bonfa, A., Busetti, F., Collarin, A., Marcomini, A., 2004. Estrogenic potential of the Venice, Italy, lagoon waters. Environ. Toxicol. Chem. 23, 1874–1880. Pojana, G., Gomiero, A., Jonkers, N., Marcomini, A., 2007. Natural and synthetic endocrine disrupting compounds (EDCs) in water, sediment and biota of a coastal lagoon. Environ. Int. 33, 929–936. Risebrough, R.W., 1976. Transfer of chlorinated biphenyls to Antarctica. Nature 264, 738–739. Robinson, B.J., Hui, J.P.M., Soo, E.C., Hellou, J., 2009. Estrogenic compounds in seawater and sediment from Halifax Harbour, Nova Scotia, Canada. Environ. Toxicol. Chem. 28, 18–25. Rodil, R., Schrader, S., Moeder, M., 2009. Non-porous membrane-assisted liquid– liquid extraction of UV filter compounds from water samples. J. Chromatogr. A 1216, 4887–4894. Routledge, E.J., Parker, J., Odum, J., Ashby, J., Sumpter, J.P., 1998. Some alkyl hydroxy benzoate preservatives (parabens) are estrogenic. Toxicol. Appl. Pharmacol. 153, 12–19. Santiago, E.C., Kwan, C.S., 2007. Endocrine-disrupting phenols in selected rivers and bays in the Philippines. Mar. Pollut. Bull. 54, 1031–1071. Sarmah, A.K., Northcott, G.L., Leusch, F.D.L., Tremblay, L.A., 2006. A survey of endocrine disrupting chemicals (EDCs) in municpal sewage and animal waste effluents in the Waikato region of New Zealand. Sci. Total Environ. 355, 135–144. Schlumpf, M., Cotton, B., Conscience, M., Haller, M., Steinmann, V., Lichtensteiger, B., 2001. In vitro and in vivo estrogenicity of UV screens. Environ. Health Perspect. 109, 239–244. Sewell, M.A., Jury, J.A., 2011. Seasonal patterns in diversity and abundance of the High Antarctic meroplankton: Plankton sampling using a Ross Sea desalination plant. Limnol. Oceanogr. 56, 1667–1681. Singley, J.E., Kirchmer, C.J., Miura, R., 1974. Analysis of Coprostanol, an Indicator of Fecal Contamination. Office of Research and Development, U.S. Environment Protection Agency, Gainesville, Florida. Szorc, C., Helfrich, S., Scianna, L.J., Panowicz, C., 2011. Seasonal Outlook for Ross Sea and McMurdo Sound 2010–2011. US National Ice Center, Suitland, Maryland, USA. Tarazona, I., Chisvert, A., Leon, Z., Salvador, A., 2010. Determination of hydroxylated benzophenone UV filters in sea water samples by dispersive liquid–liquid microextraction followed by gas chromatography–mass spectrometry. J. Chromatogr. A 1217, 4771–4778. Terasaki, M., Makino, M., Tatarazako, N., 2009. Acute toxicity of parabens and their chlorinated by-products with Daphnia magna and Vibrio fischeri bioassays. J. Appl. Toxicol., 29. Ternes, T.A., Joss, A., Siegrist, H., 2004. Scrutinizing pharmaceuticals and personal care products in wastewater treatment. Environ. Sci. Technol. 38, 392A–399A. Watts, M.M., Pascoe, D., Carroll, K., 2001;. Chronic exposure to 17a-ethinylestradiol and bisphenol-A-Effects on development and reproduction in the freshwater invertebrate Chironomus riparius (Diptera: Chrionomidae). Aquat. Toxicol. 55:, 113–124. Weber, K., Goerke, H., 2003. Persistent organic pollutants (POPs) in antarctic fish: levels, patterns, changes. Chemosphere 53, 667–678. Wu, J.-L., Lam, N.P., Martens, D., Kettrup, A., Cai, Z., 2007. Triclosan determination in water related to wastewater treatment. Talanta 72, 1650–1654. Ying, G.-G., 2006. Fate, behavior and effects of surfactants and their degradation products in the environment. Environ. Int. 32, 417–431. Ying, G.-G., Kookana, R.S., Ru, Y.-J., 2002. Occurrence and fate of hormone steroids in the environment. Environ. Int. 28, 545–551. Zenker, A., Schmutz, H., Fent, K., 2008. Simultaneous trace determination of nine organic UV-absorbing compounds (UV filters) in environmental samples. J. Chromatogr. A 1202, 64–74.

Personal care products and steroid hormones in the Antarctic coastal environment associated with two Antarctic research stations, McMurdo Station and Scott Base.

Pharmaceutical and personal care products (PPCPs) are a major source of micropollutants to the aquatic environment. Despite intense research on the fa...
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