The behaviour of di-(2-ethylhexyl) phthalate in estuaries

Marine Chemistry 68 Ž2000. 203–217 www.elsevier.nlrlocatermarchem

The behaviour of di- ž2-ethylhexyl / phthalate in estuaries A. Turner ) , M.C. Rawling Department of EnÕironmental Sciences, UniÕersity of Plymouth, Drake Circus, Plymouth PL4 8AA, UK Received 27 July 1998; received in revised form 6 July 1999; accepted 12 August 1999

Abstract Because of their extensive use as non-reactive plasticisers, phthalate esters have become widespread contaminants of the aquatic environment. There is, however, little accurate information on the solubility and sorptive behaviour of phthalates in estuaries, where contaminants generally occur at elevated concentrations and are under the influence of a number of reaction-controlling variables. In this work we have investigated the relative solubility and particle–water interactions of di-Ž2-ethylhexyl. phthalate ŽDEHP., one of the most important and abundant phthalate esters, under simulated estuarine conditions using water and sediment samples from a small, organic-rich estuary ŽBeaulieu, southern England.. Dissolved DEHP was salted out in sea water from distilled water, and a salting constant of about 1.2 l moly1 was derived. Although the compound apparently interacts with dissolved organic matter in river water, it showed no evidence of enhanced solubility over the range in DEHP concentrations studied. Adsorption onto estuarine particles was defined by the Freundlich equation, and was significantly greater in sea water than river water suggesting that the particulate organic matter is subject to either salting out or salinity-induced structural modification which improves its solvency for DEHP. Distribution coefficients Ž K D s. exhibited a strong inverse relationship with particle concentration ŽSPM., an effect defined by the equations: K D s 2.63 = 10 6 SPMy1.15 and K D s 2.64 = 10 6 SPMy0.75 ; in river water and sea water, respectively, and only partly accounted for by experimental artefacts Že.g., adsorption of DEHP to container walls.. That the gradient of the relationship was greater in river water than in sea water suggests the effect is caused, to some extent, by a particle–particle interaction mechanism Že.g., flocculation. which is inhibited at high salinities. A comparison of the results of this study with a compilation of data on DEHP distributions and partitioning in aquatic environments suggests that the estuarine behaviour of DEHP is a function of salinity, particle concentration, particulate organic carbon, and its degradation rate in the aqueous phase. A model for predicting the retention of DEHP in estuaries, incorporating these effects, is presented, and calculations indicate that, under certain hydrodynamic and chemical conditions, more than 50% DEHP discharged to a catchment may be retained by estuarine sediment, at least over a timescale equivalent to the estuarine particle residence time. q 2000 Elsevier Science B.V. All rights reserved. Keywords: phthalate esters; di-Ž2-ethylhexyl. phthalate; estuaries; adsorption; relative solubility; particle concentration effect

1. Introduction Phthalate esters have a wide variety of industrial, agricultural and domestic applications, but by far the )

Corresponding author. Tel.: q44-1752-233000; Fax: q441752-233035; E-mail: [email protected]

most important is their use as non-reactive plasticisers that improve the flexibility and workability of polymeric materials. Most phthalate esters are used to plasticise polyvinyl chloride, which contains about 35% by weight; in some products, however, the phthalate ester content may exceed 50%. Their large-scale production, which is on the order of

0304-4203r00r$ - see front matter q 2000 Elsevier Science B.V. All rights reserved. PII: S 0 3 0 4 - 4 2 0 3 Ž 9 9 . 0 0 0 7 8 - X

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A. Turner, M.C. Rawlingr Marine Chemistry 68 (2000) 203–217

several million tonnes per annum, and physical rather than chemical incorporation in the polymeric matrix, ensures that they are widespread contaminants ŽGiam et al., 1978; Wams, 1987; Thuren ´ and Larsson, 1990.. Phthalates enter the environment in wastewater effluents during the production phase, and via leaching and volatilisation from plastic products during their usage and after disposal ŽBrooke et al., 1991; Bauer and Herrmann, 1997.. They exhibit only subtle toxicity to aquatic organisms, which generally decreases with increasing alkyl chain length ŽGiam et al., 1984; DeFoe et al., 1990., and bioaccumulation through the aquatic food chain is limited by the biotransformation of phthalates, which increases with trophic level ŽStaples et al., 1997.. Present applications and emissions are, however, of concern because phthalates are suspected of mimicking estrogen ŽJobling et al., 1995; Barton and Andersen, 1998; Tyler et al., 1998.. Di-Ž2-ethylhexyl. phthalate ŽDEHP. is one of the most important phthalates, accounting for nearly 90% of European plasticiser use ŽGiam et al., 1984., and its physical, chemical and environmental properties are listed in Table 1. Reported values for solubility and octanol–water partitioning vary by several orders of magnitude, reflecting the difficulties in obtaining reliable data for phthalates. Reasons for this include the ease of sample contamination from laboratory plasticware ŽRussell and McDuffle, 1986; Leung and Giang, 1993., and experimental artefacts arising from the tendency of undissolved phthalates to form stable, homogeneous dispersions in water and surface films at the air–water interface ŽBrown et al., 1996; Staples et al., 1997.. In light of struc-

ture–activity relationships for organic chemicals, Staples et al. Ž1997. suggest values for the solubility and octanol–water partitioning of DEHP of about 3 mg ly1 and 10 7.5, respectively, attesting to a highly hydrophobic compound. Concentrations of DEHP in riverine, estuarine and marine environments are given in Table 2. Concentrations in the water column and bed sediment span four and five orders of magnitude, respectively. Some of this variability results from differences in sampling and sample pre-treatment; for example, the highest concentrations in the water column are generally reported for unfiltered samples, encompassing DEHP on suspended particles, while bed sediments exhibit a wide range of grain sizes and textures yet no consistent normalisation has been applied to the data. The remaining inter- and intra-environment variations reflect the magnitudes of local industrial and urban sources and the extent of dilution and degradation from such, and the degree of adsorption onto particles that are temporarily suspended in the water column. Given the hydrophobicity of DEHP, adsorption followed by particle deposition is likely to be an important removal mechanism from the water column in turbid environments like rivers, estuaries, and the coastal zone, which are direct recipients of waste products. Moreover, slow degradation of particle-bound forms under both aerobic and anaerobic conditions ŽJohnson et al., 1984; Roslev et al., 1998. means that bed sediment may act as both a long-term sink and secondary source of DEHP. In order to accurately predict the transport and fate of phthalates in turbid environments and improve the design and siting of effluent pipelines from

Table 1 Physical, chemical and environmental properties of di-Ž2-ethylhexyl. phthalate Property Molecular weight Melting point Boiling point Solubility @208C, Csat Octanol–water partition coefficient, log 10 K ow Vapour pressure @238C Henry’s constant Volatilisation half-life from water Chemical hydrolysis half-life @258C and pH 7 Degradation half-life in natural waters

Value

Reference y1

390.6 g mol y508C 3848C 0.6–400 mg ly1 4.20–9.64 3.4 = 10y7 –6.4 = 10y6 mmHg 1.71 = 10y5 atm-m3 moly1 15 years 2000 years 2–15 days

Boese Ž1984.; Russell and McDuffle Ž1986. Howard et al. Ž1985.; Brooke et al. Ž1990. Staples et al. Ž1997. Staples et al. Ž1997. Giam et al. Ž1984. Giam et al. Ž1984. Staples et al. Ž1997.

Table 2 Concentrations of di-Ž2-ethylhexyl. phthalate in riverine, estuarine and marine waters and sediments Water Žmg ly1 .

RiÕers and estuaries River Aire ŽUK. River Delaware Rivers Etherow and Irwell ŽUK. River Klang ŽMalaysia. River Trent ŽUK. Various rivers in central Italy Mersey Estuary ŽUK. Mississippi Estuary Lake Yssel ŽRhine Estuary. Nueces Estuary ŽTexas.

0.36–21.0 a 0.5–1.0 a 0.4–1.9 a 3.1–64.3 a 0.74–18.0 a 0.3–31.2 a 0.13–0.69 0.070 - 0.10–0.30 0.21–0.77 a

Coastal waters Chesapeake Bay Tees Bay ŽUK. Plymouth Sound ŽUK. Gulf of Mexico German Bight Tuas Bay ŽSingapore. Open Ocean North Atlantic a

Suspended sediment Žmg gy1 .

0.28–0.64

Bed sediment Žmg gy1 .

Reference

7.89–115

Long et al. Ž1998. Sheldon and Hites Ž1979. Fatoki and Vernon Ž1990. Tan Ž1995. Long et al. Ž1998. Vitali et al. Ž1997. Preston and Al-Omran Ž1989. Giam et al. Ž1978. Ritsema et al. Ž1989. Ray et al. Ž1983.

0.49–15.0 0.84–31.0 0.058–0.487 1.20 0.069

12.0–25.0 0.040–16.0

- 0.003–0.18 0.98–2.20 a 0.099–0.28 a 0.13 0.025–0.035

0.0049

0.002 0.045–0.22 0.89–2.79

Peterson and Freeman Ž1982. Law et al. Ž1991. Law et al. Ž1991. Giam et al. Ž1978. Ernst et al. Ž1988. Chee et al. Ž1996.

A. Turner, M.C. Rawlingr Marine Chemistry 68 (2000) 203–217

Location

Giam et al. Ž1978.

Unfiltered samples.

205

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the plastics industry, an understanding of the mechanisms by and extent to which phthalates interact with natural suspended particles is required. To this end, we have investigated the particle–water interactions of DEHP in river water and sea water by monitoring the uptake of a radiolabelled analogue by natural estuarine particles. The experimental design employed has proven suitable for use with trace quantities of hydrophobic organic compounds, and minimises the risk of contamination from external sources ŽRawling et al., 1998.. Samples used in the experiments have been taken from the Beaulieu, an organic-rich estuary which drains heathland and bogs of the New Forest, southern England ŽMuller, 1996; Turner et al., 1998.. Thus, there are no significant sources of phthalates in the catchment to interfere with the studies, and concentrations of dissolved organic matter are sufficiently high to enable the effects of naturally occurring organic molecules and colloids on DEHP behaviour to be evaluated.

2. Methods

Table 3 Characteristics of the filtered Beaulieu waters and sediment used in the experiments y3

Salinity, =10 Conductivity, mS pH Organic carbon, mg ly1 Cly, mg ly1 Alkalinity, mg ly1 Ca2q, mg ly1

River water

Sea water

- 0.3 210 7.25 8.97 31 50 20.8

33.80 46 000 7.96 1.41 Ž0.65a . – – –

Estuarine sediment Ž f -63 mm. @ SPM s12–1015 mg ly1 Fe, mg gy1 Ca, mg gy1 Mn, mg gy1 Organic carbon Ž foc ., % Weight loss on ignition, % Surface area, m2 gy1

8.07 b 3.27 b 430b 2.32 c 8.94 d 6.7 e

a

After exposure to a 400 W Hg discharge lamp. Available to 0.05 M hydroxylamine hydrochloride in 25% acetic acid. c Determined using a Schimadzu TOC-5000 total organic carbon analyser after digestion in 40% vrv phosphoric acid. d @5508C for 8 h. e Determined by gravimetric BET nitrogen adsorption. b

2.1. Sampling and sample preparation About 6 l riverine and marine end-member water samples, and a sample of intertidal sediment, were collected in ashed pyrex bottles with aluminium foil-lined screw caps from the Beaulieu Estuary during June 1996. Samples were foil wrapped and stored cool in transit, and on return to the laboratory, water samples were filtered through a pre-ashed 0.7 mm Whatman GFrF filter and stored in ashed pyrex bottles at 48C in the dark, while the sediment was wet sieved through a 63 mm nylon mesh. A slurry of sediment Žc. 20 g ly1 . was prepared in filtered river water and stored likewise. Characteristics of the water and sediment samples, as determined by established techniques ŽMillward et al., 1990; Rawling et al., 1998., are given in Table 3. 2.2. Experimental DEHP, uniformly labelled with 14 C and dissolved in toluene, of specific activity 3.9 = 10 8 Bq mmoly1

and radiochemical purity G 98% ŽSigma, St. Louis, MO., was used in the solubility and sorption experiments. The stock solution was diluted in n-hexane ŽRathburn HPLC grade., and an accurate quantity Žbetween 250 and 2500 Bq. of 14 C-DEHP Ž@ 37 Bq mly1 . was then dispensed into a hexane-washed glass centrifuge tube using a glass microsyringe and the solvent evaporated under a laminar flow hood. A water sample ŽMilli-Q, filtered river water or filtered sea water. of 20 ml was then pipetted into the tube, together with an appropriate volume of slurry if sorption was being investigated. The tube was stoppered and the contents then equilibrated for 16 h on a lateral shaker at 208C in the dark; earlier kinetic studies had demonstrated that this is a suitable timescale for particulate and aqueous DEHP to attain quasi-equilibrium. Sediment and water phases were separated by centrifugation at 3000 rpm for 30 min. One milliliter aliquots of supernatant and pipette rinsings were transferred to scintillation vials each

A. Turner, M.C. Rawlingr Marine Chemistry 68 (2000) 203–217

containing 4 ml Ultima Gold liquid scintillation cocktail ŽPackard Canberra., and the remaining contents of the tube were discarded. Any activity adsorbed to the centrifuge tube was then extracted twice in 6 ml n-hexane for 16 h, and a 1 ml aliquot of each extract was added to 4 ml liquid scintillation cocktail. The activities of the supernatant, pipette rinse, and tube extracts were determined using a Beckman LS6500 multi-purpose scintillation counter with factory-installed internal quench curve. The distribution coefficient, K D Žvrw., was calculated as follows: KD s

A 0 y A w y Ž AsV . As m

Ž 1.

where A s is the combined activity of the supernatant and pipette rinse ŽBq mly1 ., A 0 and A w are the activities of the original spike and the sum of tube extracts ŽBq., respectively, V is the volume of sample Žml. and m the mass of sediment Žg.. Because the particulate Žadsorbed. reactivity is derived from mass balance, accurate K D values rely on the complete recovery of the original activity added to the tube. Recoveries from spiked samples in the absence of particles were found to be dependent on a number of variables, including volume and type of solvent used to extract 14 C-DEHP adsorbed to the tube wall. The experimental conditions detailed above were, however, optimal, recovering around 80% of the original 14 C-DEHP added, and typical distributions of measured activity under such conditions are given in Table 4. It is suspected that the remaining activity escaped from the aqueous phase to the atmosphere during manipulation of unstoppered samples Žalthough loss during carrier solvent evaporation was shown to be insignificant., andror irreversibly

207

adsorbed to the centrifuge tube walls during sample incubation. Although this loss is less likely significant in the presence of particles, it should be appreciated that sorption data are subject to an unquantifiable error, an issue which is addressed in the discussion of the results. Solubility studies were conducted in Milli-Q water, Beaulieu River water and sea water, and in sea water which had been UV-irradiated for 3 h under a 400 W Hg discharge lamp to minimise the influence of dissolved organic matter. Sorption experiments in river and sea waters were conducted over a range in particle concentrations, achieved by addition of variable volumes of sediment slurry. Each experiment was undertaken in quadruplicate, and mean relative solubilities and distribution coefficients are shown with their propagated errors arising from each stage of the radiochemical analysis. The radiochemical purity of the DEHP stock solution and the solvent extracts of centrifuge tubes resulting from an experiment employing river water and 50 mg ly1 sediment was evaluated as follows. Two ml of the original Sigma product, and extracts from four tubes Žtotalling about 1000 Bq. which had been evaporated by nitrogen purge in a fume cupboard and redissolved in 2 ml hexane, were injected into a Varian 3400 gas chromatograph. Eluents were divided between a flame ionisation detector and a Lablogic radioactivity monitoring system in a ratio of 2.5:1 using a variable output splitter ŽTancell and Rhead, 1996.. The radiopurity of DEHP in both samples was found to exceed 98%, confirming the manufacturer’s stated purity and indicating neither measurable degradation of compound under the experimental conditions nor significant contamination from external sources.

Table 4 Distribution of 14 C-DEHP activity Žas mean percentage"one standard deviation Ž ns 4.. in experiments using Milli-Q water

3. Results and discussion

Association

Percentage recovery

3.1. RelatiÕe solubility

Supernatant Pipette rinse Tube extract 1 Tube extract 2 Total recovered

18"5 0.7"0.5 56"10 6.4"4.5 81"9

The aqueous concentration of DEHP, Cr , as a function of the mass added Ži.e., glass-water partitioning or relative solubility isotherm. is shown in Fig. 1 for Milli-Q water, sea water and Beaulieu

208

A. Turner, M.C. Rawlingr Marine Chemistry 68 (2000) 203–217

river water. Propagated errors generally increase with increasing amounts of DEHP present; specifically, there is a dramatic increase beyond the first three data points. This suggests that DEHP exists in true solution up to a point equivalent to its solubility, whereafter the supernatant increasingly includes colloidal dispersions of DEHP and is withdrawn through a surface film of the compound and the observed concentration results from a combination of dissolved and undissolved DEHP. Conversely, this argument provides evidence for the saturated solubility of DEHP being in the region of a few mg ly1 ŽStaples et al., 1997.. Despite the presence of undissolved DEHP, isotherms in sea water and Milli-Q water are linear throughout and the gradients of the best fit lines indicate that the relative solubility in sea water is significantly lower than in Milli-Q water. This is due to salting out of the compound, combined with any competing or additive effects resulting from the presence of marine organic matter, and may be parameterised in terms of a salting constant, s , calculated as follows ŽSchwarzenbach et al., 1993.: mq sw log Ž Csat rCsat . ss M

Ž 2.

mq sw Csat rCsat

where is the ratio of the absolute solubilities at saturation in Milli-Q and sea waters, and M is the molar salt concentration of sea water. Assuming mq sw that Csat rCsat approximates the ratio of the slopes of the relative solubility isotherms over the range in aqueous concentrations studied, m mq rmsw ŽFig. 1., and using a molar sea salt concentration of 0.5, a value for s of 1.18 l moly1 is obtained. The salting out effect over the full estuarine gradient was determined from the relative solubilities of a sub-mg spike of DEHP Žequivalent to - 3 mg ly1 . in samples of different salinity, created by mixing UV-pho-

Fig. 1. Aqueous concentration of DEHP, Cr , as a function of DEHP mass present in Ža. Milli-Q water, Žb. Beaulieu river water, and Žc. sea water. The gradients of the isotherms in Milli-Q and sea waters are derived from linear regression of all data points, forced through the origin; the gradient of the isotherm in river water is derived from linear regression of the first three data points, forced through the origin, while the best fit line through all data is an exponential function. Csat represents the best available estimate of the saturated aqueous solubility of DEHP@208C ŽStaples et al., 1997..

A. Turner, M.C. Rawlingr Marine Chemistry 68 (2000) 203–217

tolysed sea water Žminimising the concentration of marine organic matter. and Milli-Q water. Fig. 2 shows the ratio of relative solubility in Milli-Q water to relative solubility in each sample as a function of molar sea salt concentration, where the gradient of the best fit line of 1.25 l moly1 represents an alternative estimate of the salting constant. To our knowledge, there are no previously reported salting constants for DEHP from which to draw comparison. Empirical relationships have been established between s and molar volume Ž VH , cm3 moly1 . for organic compounds, and estimates of the ratio of s to VH range from 0.0018 ŽXie et al., 1997. to 0.0025 ŽAquan-Yuen et al., 1979.. A molar volume for DEHP of 474 cm3 moly1 , derived from diffusion sizes of component atoms ŽSchwarzenbach et al., 1993., yields empirical estimates for s of between 0.85 and 1.19 l moly1 , which are in reasonable agreement with the values derived from this study. Although the salting constant is an important parameter in its own right, these calculations have not accounted for the effects of riverine organic matter. Phthalates combine with natural dissolved humic material ŽCarlberg and Martinsen, 1982. and anthropogenic organic ligands ŽBauer et al., 1998. but the nature of these interactions and their effects on overall phthalate solubility have not been established. Linear regression through the first three data points and the origin of the river water solubility

isotherm, encompassing the range in aqueous concentrations typically encountered in aquatic environments ŽTable 2. and where the effects of undissolved DEHP are least significant, yields a slope of 2.12, similar to that for sea water. The presence of small amounts of dissolved salts and natural organic matter apparently has a salting out effect on DEHP such that the relative solubilities of DEHP in organic-rich river water and sea water are similar. Thus, solubility effects alone are unlikely to influence the estuarine distributions of such compounds. Unlike Milli-Q water and sea water, however, the entire relative solubility isotherm of DEHP in river water is best defined by an exponential function ŽFig. 1.. Neglecting the possible effects of undissolved DEHP at high concentrations, this suggests that complexation with natural organic matter becomes increasingly favourable with increasing amounts of DEHP, i.e., the character or hydrophobicity of the dissolved organic matter is somehow modified by its interaction with the compound. 3.2. Adsorption 3.2.1. Salinity dependence Sorption isotherms for DEHP on Beaulieu estuarine particles in river and sea waters are shown in Fig. 3. The data exhibit considerable scatter, a common problem encountered in phthalate ester adsorption studies using soils ŽRussell and McDuffle, 1986. and sediments ŽAl-Omran and Preston, 1987.. Since propagated errors increase with increasing amounts of DEHP it is suspected that undissolved DEHP may be present in dispersion or film form, although its effects on the adsorption process are unknown. The adsorption of many organic contaminants to natural particles may be defined by the Freundlich equation: P s KC n

Fig. 2. Ratio of relative solubilities of DEHP in Milli-Q water, Crmq , to saline water, CrM Žcomposed of mixtures of Milli-Q water and UV-irradiated sea water., as a function of molar concentration of sea salt, M.

209

Ž 3.

where P is the particulate or adsorbed concentration Žwrw., C is the aqueous concentration Žwrv. in the presence of particles, and K and n are constants. Linear adsorption Ž n s 1. is tested by linear regression of the data, forced through the origin, where the slope represents the estimate of K, which equates to the distribution coefficient, K D Žs PrC .. Non-linear adsorption is tested on a log–log plot, where the

A. Turner, M.C. Rawlingr Marine Chemistry 68 (2000) 203–217

210

linear fits to the adsorption isotherms, as follows ŽMeans, 1995.: log Ž K Dsw rK Drw . s sads M y log gp

Fig. 3. Adsorbed concentration, P, vs. aqueous concentration in the presence of particles, C, of DEHP in suspensions of estuarine particles Ž@150 mg ly1 . in Beaulieu river water ŽB. and sea water Ž'.. The data are defined by Eq. Ž3. for ns1 and n-1; coefficients of the regression equations are given in Table 5.

slope and intercept represent estimates of n and K, respectively, and K D varies as a function of the amount of compound present. The results of both tests to Beaulieu data are presented in Table 5 and indicate that linear fits are more significant. Estimates of either K or K D are far greater for the sea water isotherm than the river water isotherm indicating that adsorption of DEHP is more favourable in saline conditions. This effect may be parameterised in terms of an adsorption salting constant, sads , estimated using the K D values derived from the

Ž 4.

where superscripts rw and sw refer to river water and sea water, respectively, and gp is the activity coefficient of the compound in the particulate organic matter. Assuming that gp is close to unity, that is, the particulate organic matter approximates an ideal reference phase, a value of 2.2 l moly1 is obtained. This exceeds the value of the salting constant derived from the relative solubilities in Milli-Q water Žand hence also in river water. and sea water, indicating that solubility effects alone cannot account for enhanced adsorption in sea water. Some form of interaction between sea water ions and the particulate organic matter must, therefore, occur such as salting out of the organic matter, thereby improving its solvency, or salinity-induced structural changes of hydrophobic functional groups in three-dimensional space resulting in an increase in the exposure of adsorption sites ŽMeans, 1995.. 3.2.2. Particle concentration dependence Distribution coefficients for DEHP sorption to Beaulieu sediment in river water and sea water are shown as a function of particle concentration in Fig. 4 along with their corresponding values normalised

Table 5 Estimates of n and K, and the significance of the isotherm fits, p, for DEHP sorption to Beaulieu sediment in river and sea waters at a suspended particle concentration of 150 mg ly1 ŽFig. 3. Constant

River water

Sea water

n K p

1 3.77 Ž K D s 3770. 0.0055

1 49.0 Ž K D s 49 000. 0.048

n K p

0.51 20.0 0.058

0.47 115 0.13

Fig. 4. Distribution coefficients Žas log K D , and normalised with respect to particulate organic carbon, log K oc , and specific surface area, log K S . for DEHP adsorption to estuarine particles in Beaulieu river water ŽB. and sea water Ž'. as a function of particle concentration Žas log SPM .. Regression equations: river water Žsolid line., K D s 2.63=10 6 .SPMy1.15 Ž r 2 s 0.88.; sea water Žbroken line., K D s 2.64=10 6 .SPMy0.75 Ž r 2 s 0.87..

A. Turner, M.C. Rawlingr Marine Chemistry 68 (2000) 203–217

with respect to organic carbon Ž K oc ; ml gy1 . and specific surface area Ž K S ; ml my2 .. In these experiments, it was ensured that the aqueous concentration did not exceed about 2 mg ly1 , minimising possible effects caused by undissolved compound as dispersion or film. The best fit lines through the data are of the form: log K D s yblog SPMq log a

Ž 5a .

or: K D s aSPMyb

Ž 5b .

where SPM is the particle concentration, and a and b are constants whose values in river water and sea water are 2.63 = 10 6 and 1.15, and 2.64 = 10 6 and 0.75, respectively. Deviation from a linear two-phase equilibrium, commonly referred to as a particle concentration effect ŽPCE., has been reported for a number of hydrophobic organic pollutants in laboratory studies, although its exact causeŽs. has not been unequivocally resolved ŽMcKinley and Jenne, 1991.. The PCE observed herein may be partially accounted for by loss of compound to the glassware, an artefact which results in an overestimation of K D in proportion to the reciprocal of particle concentration. Recalculating K D s in Beaulieu river water, assuming a worst case scenario of 20% loss of compound ŽTable 4. which is independent of the amount of particles present, yields values of a and b of 2.42 = 10 6 and 1.22, respectively, which are not dissimilar to those derived from the mass balance approach ŽEq. Ž1... Many authors attribute a PCE to the inadequate sediment–water phase separation achieved by centrifugation ŽGschwend and Wu, 1985; Koelmans and Lijklema, 1992., while alternative explanations are based on the effects of particle interactions ŽMackay and Powers, 1987; Servos and Muir, 1989.. Flocculation of colloidal material and fine particles occurs in the presence of dissolved ions which neutralise the surface particle charge, the extent of flocculation being dependent on particle collision frequency and the nature or ‘stickiness’ of the organic matter coating the particles. An increase in aggregation is, therefore, predicted with increasing particle concentration, resulting in a reduction in surface area or exposed organic matter available for adsorption. Laboratory experiments employing clay mineral suspensions and field measurements of particle size

211

distributions in estuaries suggest that flocculation does not continue above salinities of a few g ly1 and that deflocculation may in fact occur in sea water ŽEisma, 1993.. These observations are compatible with the data presented here in that the estimate of b, or the sensitivity of adsorption to particle concentration, is reduced in saline water. Presumably, the ‘stickiness’ of the surface organic matter, which may be some function of its orientation in three dimensional space ŽMurphy et al., 1994., is reduced by its interaction with sea water ions. This effect may be parameterised in terms of the variation of the adsorption salting constant as a function of particle concentration as follows. Assuming that the particulate organic matter approximates an ideal reference phase, Eq. Ž4. can be rewritten, thus:

sads s

log Ž K Dsw rK Drw .

Ž 6.

M

or, substituting the PCE ŽEq. Ž5b.. for the distribution coefficients in river water and sea water:

sads s

log Ž w aSPMyb x sw r w aSPMyb x rw . M

Ž 7.

With respect to DEHP in the Beaulieu, Eq. Ž7. reduces to:

sads s

log Ž SPM 0.4 . M

Ž 8.

since the estimates of a in river water and sea water are almost identical, and where 0.4 is the numerical difference between the estimates of b in river water and sea water. Thus, the onset of the effects of salting out on DEHP adsorption is predicted at a particle concentration of 1 mg ly1 ; below this concentration, ‘salting in’ is predicted, although in experiments conducted in the absence of particles ŽFig. 1. no salting effects were observed. Although it was not an objective of this study to further our mechanistic understanding of the PCE, the results have, nevertheless, demonstrated its significance for a highly important contaminant. Moreover, that a PCE is reported in the field for many organic contaminants ŽEadie et al., 1990; Bergen et al., 1993. suggests it is the result of some environmental Žpossibly particle interactive. process which is replicated under laboratory conditions, and as

212

A. Turner, M.C. Rawlingr Marine Chemistry 68 (2000) 203–217

such, we might reasonably expect estuarine distributions of DEHP to be subject to the same control. In addition, experiments conducted in this study have shown that sorption of DEHP to estuarine particles is dependent on salinity and, presumably, particle character Žspecifically, the fraction of organic carbon, f oc .. In order to explore the implications of these results and the general estuarine behaviour of DEHP, Beaulieu data are next compared with sorption data and field data derived from a number of other studies. 3.3. Factors controlling the estuarine distributions of DEHP Table 6 presents a summary of field- and laboratory-derived K D s for DEHP in rivers, estuaries and coastal waters, with accompanying water and sediment characteristics, where available. The data are plotted as a function of particle concentration in Fig. 5, along with their corresponding salinity values and the Beaulieu regressions. Sorption data for the Humber Estuary were derived by monitoring the uptake of 14 C-DEHP in batch mixing experiments using unfiltered river and marine end-member samples ŽZhou and Rowland, 1997.. Although the data may be defined by Eqs. Ž5a. and Ž5b., the estimates of a Žs 2.64 = 10 17 . and b Žs 5.13. are inflated by the large range in salinity, relative to particle concentra-

Fig. 5. Distribution coefficients Žas log K D . for DEHP as a function of particle concentration Žas log SPM. in various rivers, estuaries and coastal waters. Sources of information and samplingrexperimental approaches adopted are given in Table 6. Salinity data are shown in parentheses, and Beaulieu regressions are delineated. Open symbols represent field data and closed symbols represent laboratory-derived data.

tion, employed in the experiments. Sorption data for the Mersey Estuary, derived from field measurements of particulate and dissolved DEHP throughout the estuary ŽPreston and Al-Omran, 1989., exhibit an

Table 6 Distribution coefficients for di-Ž2-ethylhexyl. phthalate in rivers, estuaries and coastal waters Location Laboratory incubations Unfiltered Humber Estuary water Unfiltered Humber marine water Marine bed sediment in filtered sea water Mersey Estuary bed sediment in filtered sea water R. Mississippi bed sediment in artificial river water Field measurements Mersey Estuary Lake Yssel ŽRhine. R. St. Lawrence R. Niagara a b

Salinity Ž=10y3 .

SPM Žmg ly1 .

f oc Ž%.

log K D Žml gy1 .

Reference

0.4 23.8 ; 35

302 254 920–6900

5.71 9.81 -1

4.75 5.11 3.71

Zhou and Rowland Ž1997. Sullivan et al. Ž1982.

; 35

1000

1.06

2.52

Al-Omran and Preston Ž1987.

1.6 a

2000–20000

0.15–1.88

2.65–3.77

Williams et al. Ž1995.

13.6–28.9 ; 0.5 b - 0.1 - 0.1

940–2670 ; 100 b 3 7.8

– 6.7–45.8 – –

2.88–3.61 4.79–5.40 5.58 4.67

Preston and Al-Omran Ž1989. Ritsema et al. Ž1989. Staples et al. Ž1997. Staples et al. Ž1997.

0.01 M CaŽNO 3 . 2 . Typical salinities and particle concentrations encountered in the Lake ŽSalomons et al., 1981..

A. Turner, M.C. Rawlingr Marine Chemistry 68 (2000) 203–217

increase in K D with particle concentration. Here, the effects of particle concentration and salinity are apparently more than compensated for by variations in particle character; although particulate organic carbon was not determined, twenty five-fold differences in particulate lipid content are reported. Pooled K D s span three orders of magnitude and there is a general reduction in K D with increasing particle concentration. Inter-environment variability reflects differences in water and particle characteristics, as well as sampling Že.g., bed sediment vs. suspended sediment., analytical, and experimental differences. Normalisation of the pooled data with respect to f oc , where reported or calculable, gave rise to more scatter, resulting, in part, from inconsistent approaches used to assess organic carbon content Žloss on ignition, elemental analysis, lipid extraction., but also reflecting differences in the characteristics Že.g., polarity, age. of organic coatings on estuarine particles. The compiled data demonstrate, therefore, that the partitioning of DEHP in estuaries is a complex function of salinity, particle concentration and particulate organic content. Accurate parameterisation of each effect, in terms of empirical algorithms such as Eqs. Ž4., Ž5a. and Ž5b., is required for pollution modelling and impact evaluation purposes, and will rely on further site-specific, process-orientated studies. Additionally, field distributions of DEHP are affected by its biodegradation ŽTable 1., an effect not mimicked in laboratory studies which rely on the addition of some microbial inhibitor ŽSullivan et al., 1982. or termination of the experiments within a few days before DEHP has degraded significantly ŽWilliams et al., 1995; this work.. The microbial degradation half-life of DEHP in river water is on the order of several days to several weeks ŽGiam et al., 1984; Ritsema et al., 1989; Staples et al., 1997., the precise time being dependent on temperature, dissolved oxygen concentration and bacterial density. This timescale encompasses the hydrodynamic flushing times of many estuaries ŽMillward and Turner, 1995. and degradation is, therefore, likely to have a significant effect on the concentrations and distributions of dissolved DEHP in the estuarine water column, with minimum concentrations predicted during summer months when temperatures, flushing times and autotrophic activity are greatest.

213

Degradation of phthalates adsorbed to sediments is slower than that of their aqueous counterparts. For example, Johnson et al. Ž1984. found that - 2% DEHP had degraded in aerobic, fresh water sediments incubated at 228C for 14 days. However, correlation between plasticiser usage and phthalate concentrations in sediment cores ŽPeterson and Freeman, 1982; Giam et al., 1984. suggests that in situ degradation may be considerably slower than this. It is predicted, therefore, that K D s derived in the field are greater than those derived in the laboratory under identical conditions. That Mersey data for DEHP derived under laboratory conditions in the presence of a bacterial inhibitor are lower than field K D s determined at equivalent salinities and particle concentrations ŽFig. 5. lends support to this hypothesis. 3.4. The retention of DEHP in estuaries An important consideration for water quality managers is the capacity of an estuary to filter or retain contaminants ŽMeade et al., 1990; Brunk et al., 1997; Turner et al., 1999.. Having identified the factors which control the behaviour of DEHP a model for its retention in estuaries is now presented. If it is assumed that degradation in the sediment is negligible in the short-term Že.g., compared with estuarine particle residence times., then the retention of DEHP will be governed by two competing effects. Firstly, the degree of adsorption onto suspended particles, which is dependent on the suspended particle concentration and distribution coefficient, K D , and secondly, the rate of decay in the aqueous phase, which is dependent on the temperature and a number of water quality parameters Ždissolved oxygen, microbial density.. In practice, therefore, the retention is dependent on the quality and transit time of water from the point of discharge to a region of high turbidity Že.g., the turbidity maximum zone, TMZ., and the extent of bed sediment resuspension or suspended particle trapping in this region. The fraction adsorbed in the TMZ, f p , is calculated as follows: 1 fp s 1 y Ž 9. 1 q K D SPMr10 6 where the term on the right hand side defines the aqueous Žnon-adsorbed. fraction, and SPM is the concentration of particles Žmg ly1 . resuspended from

A. Turner, M.C. Rawlingr Marine Chemistry 68 (2000) 203–217

214

the bed andror trapped at the head of the intruding saline undercurrent. Incorporating a particle concentration effect ŽEqs. Ž5a. and Ž5b.. yields: fp s 1 y

1 1 q Ž aSPMyŽ by1. . r10 6

Ž 10 .

where the constants a and b are environment-specific and salinity dependent. This model assumes a continual supply of uncontaminated particles to the TMZ, a condition which is approximated in macrotidal estuaries by upestuary, tidal transport of marine particles ŽAvoine, 1986.. In reality, however, Eqs. Ž9. and

Ž10. require empirical modification to account for the extent of pre-existing particulate contamination or the sorption capacity of suspended particles for DEHP. The fraction of the total DEHP discharged from a point source upriver which is retained in the estuary, f r , is calculated as follows: f r s f p exp Ž ykT .

Ž 11 .

assuming first-order degradation of the compound, where T is the water transit time from source to the TMZ and k is the decay constant which is dependent

Fig. 6. Estuarine retention of DEHP, f r , as a function of suspended particle concentration, SPM, calculated according to Eqs. Ž10. and Ž11. using Beaulieu regression data ŽFig. 4., for different water transit times, T, and for two DEHP degradation constants ŽTable 1..

A. Turner, M.C. Rawlingr Marine Chemistry 68 (2000) 203–217

on water quality and temperature. Estimates of decay half-lives for DEHP in natural waters under realistic conditions of temperature and water quality range from 2 to 15 days ŽTable 1., equivalent to a range in decay constants of 0.347 to 0.0462 dayy1 . Eq. Ž11. also assumes that transport of DEHP to the TMZ is solely in the aqueous phase; for particle-water partitioning upriver of the TMZ, estuarine retention is augmented Žor reduced. by an amount dependent on the extent of Ži. adsorption to riverine particles, and Žii. retention Žor escapement. of these particles by the estuarine circulation. Fig. 6 shows the retention of DEHP, calculated according to Eqs. Ž10. and Ž11., as a function of suspended particle concentration Žwithin the range 1 to 1000 mg ly1 . for different water transit times and the extreme decay constants; values for the constants a and b were based on Beaulieu data in river water and sea water ŽFig. 4.. When the TMZ is located towards the river end-member Ž a s 2.63 = 10 6 ; b s 1.15. an increase in particle concentration is more than countered by the concomitant reduction in K D Ža condition which applies when b ) 1.0. and a reduction in retention with increasing particle concentration arises. Under these conditions, it is predicted that more than half the DEHP which is discharged into a river may be retained in the estuary, depending on particle concentration, when the transit time is less than a day Ž k s 0.347 dayy1 . or less than about 8 days Ž k s 0.0462 dayy1 .. When sorption occurs in the marine reaches of the estuary Ž a s 2.64 = 10 6 ; b s 0.75. the overall retention is greater and an increase with increasing particle concentration occurs Ža condition which applies when b - 1.0.. Such results are environment-specific and indicative rather than absolute. Assumptions of the model which require validation or further investigation include negligible DEHP degradation in the bed sediment ŽGiam et al., 1984., and instantaneous DEHP sorption which is independent of the quality of particles Že.g., nature of organic carbon and degree of contamination. comprising the TMZ. Qualitatively, however, the model results indicate important seasonal differences in the capacity of estuaries to retain hydrophobic micropolluants which are subject to degradation. Thus, retention is greatest during winter when autotrophic activity, temperatures and water

215

transit times are minimal, resuspension of bed sediment is greatest, and the TMZ is propagated seaward by riverine currents. 4. Conclusions In this study, the relative solubility and sorptive behaviour of 14 C-labelled di-Ž2-ethylhexyl. phthalate ŽDEHP. have been determined in laboratory experiments employing end-member water samples and intertidal sediment from an organic-rich estuary ŽBeaulieu, UK.. Particle-water interactions were found to be highly complex inasmuch as Ži. relative solubility in the presence of riverine organic matter was nonlinear, and Žii. partitioning exhibited a strong dependence on salinity and particle concentration. The former observation may be attributed to the existence of some DEHP in undissolved form under the experimental conditions, andror synergistic interaction of DEHP with dissolved organic matter. The latter observations suggest that the solvency of the particulate organic matter for DEHP is improved by its interaction with sea water ions, and that the sorption capacity of estuarine particles is reduced through some, as yet unresolved, particle–particle interaction mechanism. An examination of published data on DEHP concentrations and partitioning in estuaries suggests that distributions are controlled, additionally, by microbial degradation in the aqueous phase and the concentration and quality of particulate organic matter. Incorporation of partitioning data and degradation rates for DEHP into a simple model indicates that, under realistic environmental conditions, more than 50% DEHP discharged to a catchment may be retained in the estuary, at least with respect to a timescale equivalent to the estuarine particle residence time. Better prediction of the transport and fate of DEHP and other phthalate esters in estuaries will require accurate, site-specific parameterisation of sorption and degradation effects and their incorporation into a suitable hydrodynamical modelling framework. Acknowledgements The authors are grateful for the financial support from the Natural Environment Research Council and

216

A. Turner, M.C. Rawlingr Marine Chemistry 68 (2000) 203–217

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