evaluation of organic contamination in urban groundwater surrounding...

Evaluation of organic contamination in urban groundwatersurrounding a municipal landfill, Zhoukou, China

D. M. Han & X. X. Tong & M. G. Jin &

Emily Hepburn & C. S. Tong & X. F. Song

Received: 9 February 2012 /Accepted: 23 July 2012 /Published online: 8 August 2012# Springer Science+Business Media B.V. 2012

Abstract This paper investigates the organic pollutionstatus of shallow aquifer sediments and groundwateraround Zhoukou landfill. Chlorinated aliphatic hydro-carbons, monocylic aromatic hydrocarbons, halogenat-ed aromatic hydrocarbons, organochlorine pesticidesand other pesticides, and polycyclic aromatic hydrocar-bons (PAHs) have been detected in some water samples.Among the detected eleven PAHs, phenanthrene, fluo-rine, and fluoranthene are the three dominant in most ofthe groundwater samples. Analysis of groundwater sam-ples around the landfill revealed concentrations of PAHs

ranging from not detected to 2.19 μg/L. The resultsshow that sediments below the waste dump were lowin pollution, and the shallow aquifer, at a depth of 18–30 m, was heavily contaminated, particularly during thewet season. An oval-shaped pollution halo has formed,spanning 3 km fromwest to east and 2 km from south tonorth, and mainly occurs in groundwater depths of 2–4 m. For PAH source identification, both diagnosticratios of selected PAHs and principal component analy-sis were studied, suggesting mixed sources of pyro- andpetrogenic derived PAHs in the Zhoukou landfill.Groundwater table fluctuations play an important rolein the distribution of organic pollutants within the shal-low aquifer. A conceptual model of leachate migrationin the Quaternary aquifers surrounding the Zhoukoulandfill has been developed to describe the contamina-tion processes based on the major contaminant (PAHs).The groundwater zone contaminated by leachate hasbeen identified surrounding the landfill.

Keywords Landfill . Organic contamination .

Hydrogeology . PAHs . Conceptual model

Introduction

Groundwater, used mostly for irrigation, drinking water,and municipal water supplies, is essential to the econom-ic viability and livability of many cities in China.Groundwater contamination caused by human activitiesis universal, with extensive pollutant sources such as

Environ Monit Assess (2013) 185:3413–3444DOI 10.1007/s10661-012-2801-z

D. M. Han :X. F. SongKey Laboratory of Water Cycle & Related Land SurfaceProcesses, Institute of Geographic Sciences and NaturalResources Research, Chinese Academy of Sciences,Beijing 100101, China

D. M. Hane-mail: [email protected]

X. X. Tong (*)School of Water Resources & Environment,China University of Geosciences,Beijing 100083, Chinae-mail: [email protected]

M. G. Jin :C. S. TongSchool of Environmental Studies,China University of Geosciences,430074 Wuhan, China

E. HepburnSchool of Earth Sciences, The University of Melbourne,Parkville 3010, Australia

wastewater, landfill leachate, storage and disclosure ofpetroleum products, and pesticide and herbicide use.Unsuitable disposal of organic products can result inunreasonable emissions and harmful byproducts enteringthe geological environment, causing groundwater con-tamination. The burial of municipal solid waste in land-fills is the most common disposal alternative in mostcountries. According to the investigation in Henan partof Huaihe River Basin, the total amount of domesticgarbage from 17 primary cities is up to 3,640×103 m3,with approximately 80 landfill sites in major towns(Tong 2012). Most of early established landfills do nothave an engineered liner, leachate collection system, orengineered cover system. Thus, landfill leachate couldhave the potential to pollute soil andwater system directly.The current policies surrounding landfill management inChina are guided by the Standard for Pollution Control onthe Landfill Site of Municipal Solid Waste (GB 16889-2008), which is released by Ministry of EnvironmentalProtection of the People’s Republic of China.

Organic contamination issues resulting from landfillin many urban areas are of particular concern to localauthorities and scientists, since agricultural activities arecarried out close to near these cities and since ground-water is a major supply of both irrigation and domesticwater. Most volatile halogenated compounds, even atlow concentrations, are probably carcinogens or muta-gens (Baudoin et al. 2002). These have been paid inter-national attention and are strictly controlled by waterand air quality standards. Substantial research on theenvironment, such as polycyclic aromatic hydrocarbons(PAHs), organochlorine pesticides (OCPs), and tetra-chloroethylene (Nielsen et al. 1995; Persson et al.2006; Eggen et al. 2010), has theoretical significanceand applicability to field based studies of landfills suchas those located throughout the numerous small urbancenters in China (Zhou and Maskaoui 2003). Due toPAHs ubiquitous occurrence, recalcitrance, bioaccumu-lation potential, and carcinogenic activity, the PAHshave gathered significant environmental concern.PAHs have a detrimental effect on the flora and faunaof affected habitats, resulting in the uptake and accumu-lation of toxic chemicals in food chains, which causeserious health problems and/or genetic defects inhumans. The major potential environmental impactsrelated to landfill leachate are pollution of groundwaterand surface water. The leaking of strongly reducedlandfill leachate, high in organic matter, into a shallow,presumably aerobic aquifer creates a very complicated

environment owing to redox processes, biodegradation,dissolution/precipitation, complexation, ion exchange,and sorption processes (Christensen 1992).

This research focuses on one municipal landfill inZhoukou city, which is located at the Huaihe RiverBasin (Fig. 1). The results of an investigation into sevenbig rivers in China in 1993 showed that the main pollu-tion type in Huaihe River was organic contamination(Cui and Fu 1998). Huaihe River has the highest popu-lation density of these rivers and, with rapid develop-ment of economic society, gradually increasing waterdemands will conflict with water shortages in manycities, including Zhoukou city. Groundwater is the mainsource of water supply for industrial, agricultural, and,locally, domestic water in Zhoukou city, and it is facinggroundwater quality problem. There is a major concernthat urban pollution can affect the production wells inthis and other similar settings in Northern China. It istherefore both necessary and urgent to develop reason-able groundwater utilization practices and effective pro-tection of the resource. If no successful measures aretaken for reducing the leakage and transport of pollu-tants in urban soils and groundwater, the accumulationof contaminants can not only degrade soil quality butalso pose a health risk to humans and the ecosystem.This study investigated the hydrogeological conditionsand the status of organic pollution in groundwater andsoil in a typical landfill around Zhoukou city. Theobjectives of this study were therefore (1) to investigatethe extent of chlorinated aliphatic hydrocarbons(CAHs), monocylic aromatic hydrocarbons (MAHs),halogenated aromatic hydrocarbons (HAHs), OCPsand other pesticides, and polycyclic aromatic hydrocar-bons (PAHs) in aquifer sediments and surface andgroundwater around Zhoukou landfill, and (2) to deter-mine the potential sources and pathways of PAHs as themain contaminant in polluted groundwater. Finally,some suggestions are put forward for the reasonabledevelopment of water resources and better protectionof the eco-geological environment. These can provide abasic framework and scientific foundation for protectingand managing groundwater resources.

Study area description

Landfill background

The landfill is located in the north of Zhoukou city inHenan province, China. The study area lies between

3414 Environ Monit Assess (2013) 185:3413–3444

longitudes 114°36.7′ and 114°40.8′ E and latitudes 33°37.1′ and 33°39.8′ N, and has an area of 18 km2. It isan alluvial depositional plain, bordered by LuDongTrunk Canal to the north, Ying River to the south,Jialu River to the west and the Low-lying Gully to theeast (Fig. 1). Elevation ranges between 45 and 51 ma.s.l. with land surface gradient ranging from 1/3,000to 1/6,000. The area has a continental monsoonalclimate, with an annual mean air temperature of14.6 °C, a mean rainfall of 790.8 mm (averaged be-tween 1951 and 2004) and a mean potential evapora-tion of 1,736 mm. As much as 54 % of yearlyprecipitation is concentrated in July, August, andSeptember. The main rivers flowing through the studyarea include Ying River and Jialu River, which areboth perennial rivers and belong to the Huaihe Rivernetwork. The Ying River and Jialu River are

characterized by small bed slope and big changes ofwater table and flux. During flood season, surfacewater recharges local groundwater. After the conflu-ence of Ying River and Jialu River, the stream flowstowards the southeast and into Huaihe River. YingRiver and Jialu River converge at the south ofZhoukou city and flow towards the southeast. Onlyduring the devastating floods, the river waterrecharges groundwater, and river receives groundwa-ter discharge in most cases.

Landfilling operations at the Zhoukou site spanneda period of 13 years, from 1998 to 2010. The landfillsite was the borrow pits of the Beijiao brickworksbefore 1998. The activities for excavating soil at theformer brickworks have resulted in the formation ofpits and trenches with different sizes and depths. In1998, the brickworks were closed and began piling up

Fig. 1 Map of the sampling sites around the Zhoukou landfill,China. 1 residential area, 2 landfill, 3 farmland, 4 orchard, 5waters, 6 groundwater monitoring wells (sampling wells withlabels in Table 1), 7 surface water sampling sites, 8 shallow

groundwater table contours (m.a.s.l), 9 major groundwater-flowdirection, 10 flow direction of surface water, 11 sewage ditch, 12local factory. The dashed line delineates the pollution range.Groundwater contours from December 2009

Environ Monit Assess (2013) 185:3413–3444 3415

household and construction waste, ceasing in 2010.The length of the landfill site measures 200 m fromsouth to north and 90 m from east to west and iscurrently surrounded by the city planning area. Thedepth of the solid waste disposal ranges from 6 to 9 m.The waste generated by city life has been stacked inthe landfill since 1998, and the landfill was closed in2010. The accumulative amount of municipal solidwaste, mainly composed of household waste, is nowup to approximately 140×103 m3, including garbage,trash, and septic tank waste, derived from houses,apartments, hotels, campgrounds, and picnic grounds(Tong 2012). The construction of this landfill has nodesign features intended to prevent movement ofleachate into the ground water. The landfill does nothave an engineered liner, leachate collection system,or engineered cover system. There is a wastewaterdischarge canal to the east of the landfill. The landfillleachate and wastewater discharge have resulted inserious contamination to the ambient groundwaterand surface water. This landfill with a life of over10 years was managed by the Zhoukou CityEnvironmental Sanitation Management Office.

Hydrogeological setting

The hydrogeological conditions play an important rolein controlling the distribution of groundwater organiccontamination. Zhoukou city is located in the southernpart of the Yellow River alluvial–diluvial fan. Thevadose zone is characterized by coarse grains.Generally, there are two aquifers within 55 m deptharound the Zhoukou landfill. Figure 2 shows the char-acterization of the hydrogeology based on drillingaround the landfill. The shallow aquifer at the depthof 11–25 m, mainly consists of fine sand and siltysand, is the major aquifer exploited for local irrigation.The thickness of the aquitard on top of the first aquiferis approximately 11 m. The deep aquifer consists ofinterbedded fine sand and silty sand is distributed atthe depth of 45–52 m, with a hydraulic conductivity of12–16 m/day (Qu 2010). One weak permeable layercomposed of silt and silty clay with 20 m thickness isdistributed between the two aquifers. Aquifer sandthickness becomes slightly thinner towards the YingRiver. From pumping testing at the field site (Tong2012), the hydraulic conductivity was determined to

Fig. 2 Simplified hydrogeological sections of the Zhoukoulandfill. ZKE, ZKW, ZKN, and ZKS are the groundwater mon-itoring wells. ZKC is located at the center of the landfill. ZKS′ is

located at some 150 m south of the landfill. The hydraulicparameters of the aquifers are obtained from the results of thepumping test (Tong 2012)


be in the range 1.2–63.4 m/day. The main rechargesources of the shallow aquifer include vertical precip-itation infiltration, lateral recharge by rivers andcanals, and irrigation return flows. The groundwaterdischarge also includes human exploitation and drain-age to rivers. Local groundwater flow under naturalconditions is towards the southeast with a gradient of1/3,000 to 1/5,000.

One control bore (17 m depth) has been drilled inthe center of the landfill. The garbage layer, composedof domestic waste material, is located between 0 and9.3 m. The strata intersected by the drilled bore can beseen in Fig. 2. According to drilling data, the depth ofwaste dump at the center of the landfill reached 9.3 m.There is a silty clay layer of 2–3 m thickness below thedepth of 9.4–13.8 m, which has a certain protectivefunction for the shallow aquifer. However, the silt andfine sand below the silty clay layer is less protective(more permeable). In the east, west, south, and northof the landfill, four bores, namely, ZKE, ZKW, ZKS,and ZKN, respectively, have been drilled at a depth of27 m in order to monitor the groundwater table andsample for organic pollutants. The mean groundwatertable depth around the landfill was 3.6 m in May 2009and 3.9 m in December 2009.

Some criss-crossing ditches and scattered ponds wereconstructed due to the irrigation and drainage needs of theregion. These ditches, channels, pits, and ponds can storeprecipitation and surface water during the rainy seasonand become receivers of wastewater emission in the dryseason; this is one of the sources of potential pollution ofshallow groundwater. The N–S drainage ditch passesthrough the eastern landfill and probably provides localrecharge to groundwater. Artificial exploitationmakes thesurrounding water level slightly lower than that in theposition of nonexploitation areas. A water-table moundexists beneath the landfill in response to the rainfallinfiltration, and this has diffused into the surroundings.

Different types of pollution, such as waste from thelocal winery, food factory, pesticide factory, gas station,municipal landfill, and garbage pollution treatment plant(locations in Fig. 1), may be potential pollution sourcesto the shallow aquifer. Pollution sources may also in-clude sewage, municipal wastewater, and agriculturalfertilizers. Previous investigation (Tong 2012) showsthat groundwater pollution away from urban areas andvillages are relatively mild in this area; hence, the majorsource of groundwater pollution is likely to be themunicipal landfill and possibly sewage ditches, whichreceive wastewater discharge.

Fig. 3 Sediment distribution in the core profile located at the center of the landfill, and the main organic pollutants’ concentrations inthe sediment samples. The water depth is about 3.5 m


Material and methods

Sample collection

Field investigations were carried out around the land-fill site in the north of Zhoukou city.

Sediment sampling in the shallow aquifer

Seven sediment samples from different depths werecollected by drilling the control bore in May 2009.The bore was located at the center of the landfill,where depth to water was 3.5 m. Sediment sampledepths were 9.3, 9.3–9.5, 9.7–9.9, 10.1–10.3, 10.5–10.7, 10.9–11.1, and 11.3–11.5 m (Fig. 3). Within3 min, 5 g of sediment was collected in an amberbottle (40 mL), with 5 mL NaHSO4 (20 %) and sub-jected to one magnetic stirring, which was ultrasoni-cally cleaned by methanol in advance. The collectedsamples were sealed tightly, placed upside down slow-ly and stored at 0–4 °C in the freezer. Sample analysis

began as soon as possible after returning to the labo-ratory. The measured results are shown in Table 1.

Water sampling

During three main sampling campaigns in December2008, May 2009, and December 2009, 40 wells and 10surface water sites were investigated for organic mat-ter analysis (Table 2).The locations of the samplingstations are shown in Fig. 1. The wells, includingproduction and observation wells, were purged beforesampling, and groundwater was sampled by pumpingafter constant values of conductivity and redox poten-tial had been established. Most selected samplingpoints for groundwater were situated near the landfill.Seven sampling sites were selected for surface water(sampling depth 0.5 m below surface) (Fig. 1). A freshsample tube was used for each piezometer to preventcross contamination. All samples were filtered throughGF/Fs (Whatman, Brentford, UK) to separate the par-ticulate from the dissolved fraction. The samples were

Table 1 Concentrations of organic compounds in the sediment samples in the Zhoukou landfill

Compound name Sampling depth (m)

9.3 9.3–9.5 9.7–9.9 10.1–10.3 10.5–10.7 10.9–11.1 11.3–11.5

Phenol-D5 1.54 1.59 1.30 1.83 1.80 0.71 1.37

Phenol,2-fluoro 1.53 1.62 1.34 1.74 1.73 0.66 1.42

Phenol 0.06 – – – 0.05 – –

Phenol,3-methyl 0.00 – – – – – –

Nitrobenzene-D5 1.36 1.48 1.11 1.78 1.93 0.69 1.26

Naphthalenea 0.06 – – 0.03 0.04 – 0.04

Naphthalene,1-methyla 0.06 – – – – – –

1,1-Biphenyl ss 1.41 1.42 1.16 1.51 1.69 0.69 1.27

Dibenzofurana 0.04 0.03 – – 0.03 – –

Fluorenea 0.04 – – – – – –

Phenanthrenea 0.11 – – – – – –

Anthacenea 0.14 – – – – – –

Dibutyl phthalate 0.00 – – – 0.16 – –

p-Terphenyl-d14 2.05 1.82 1.57 1.85 2.44 0.87 1.23

1,2-Benzenedi acid ,disooctyl ester 1.42 – – – – – –

Benzo(b)fluoranthenea 0.54 – – – – – –

∑All compounds 10.35 7.97 6.48 8.75 9.87 3.77 6.58

∑PAHs 0.99 0.03 0.00 0.03 0.07 0.11 0.04

Units are in nanograms per milligram

(–) “not detected”a PAHs


Table 2 Physico-chemical values of water samples around the landfill

Locationsite

Samplingtime

Well depth(m)

Water depth(m)

Utilization T (°C) pH Turbidity EC(μS/cm)

Eh (mV) DO(mg/L)

ZA December 2008 30 3.4 Agricultural irrigation 16.6 7.7 0.4 964 256 6.38

ZB December 2008 50 3.5 Domestic water 14.6 7.7 0.9 1,086 195 5.46

ZC December 2008 SU Sewage water 19.6 7.7 90.1 1,602 −208 4.03

ZD December 2008 SU water from pond 13.9 8.1 82.5 638 138 3.06

ZE December 2008 SU Waste leachate 11.1 7.9 79.2 3,255 −184 5.76

ZF December 2008 18 3.3 Agricultural irrigation 16.8 7.3 0.5 3,365 328 1.54

ZG December 2008 20 3.3 Domestic water 17.3 7.3 0.7 2,235 283 4.97

Z6A December 2008 16 3.0 Domestic water 17 7.1 0.3 1,734 149 5.59

Z6 December 2008 20 2.6 Domestic water 17 7.3 0.2 1,770 160 8.67

SW29 December 2008 9 4.0 Domestic water 16.3 7.5 0.1 1,255 60 6.1

SW49 December 2008 14 5.0 Domestic water 18.3 7.0 1.9 1,231 137 0.48

SW59 December 2008 6 3.0 Domestic water 17.5 7.5 1.6 1,325 −5 0.54

ZK1 December 2008 300 Urban water supply 22 8.5 0.1 893 72 2.41

ZKE May 2009 27 3.4 Observation well

ZKW May 2009 27 3.6 Observation well 18.3 8.9 3.5 396 98 1.07

ZKS May 2009 27 3.5 Observation well 18.4 7.5 10.3 961 3 1.98

ZKN May 2009 27 4.0 Observation well 19.3 6.9 17.2 3,644 115 1.92

ZF May 2009 18 4.1 Agricultural irrigation 17.6 7.6 0.9 1,316 151 2.55

ZG7 May 2009 9 Observation well


ZG10 May 2009 9 2.0 Observation well 18.9 7.5 9.6 1,076 73 2.29

ZG11 May 2009 9 3.7 Observation well 18.7 6.8 4.6 3,430 25 2.29


Z16 May 2009 30 4.3 Agricultural irrigation

Z34 May 2009 28 3.9 Agricultural irrigation 17.6 7.6 4.3 1,809 194 2.35

Z40 May 2009 30 4.1 Agricultural irrigation 17.2 7.6 4.1 1,162 189 1.98

DW09 May 2009 16 3.9 Domestic water 18.9 7.4 4.8 1369 153 1.85

DW18 May 2009 30 4.1 Domestic water 21.5 8.0 1.8 2,930 106 2.05



ZC May 2009 SU Sewage water

ZD May 2009 SU Water from pond

SUJL May 2009 SU River water 22.7 7.8

SULG May 2009 SU Sewage water 22.4 7.9

SUY May 2009 SU River water 18.9 7.9

SULD May 2009 SU Agricultural irrigation 20.5 7.8

ZKE December 2009 27 4.0 Observation well 17.8 7.3 1,676 −91.3ZKW December 2009 27 3.6 Observation well 16.6 7.2 1,652

ZKS December 2009 27 4.1 Observation well 17.5 7.4 1,353 −87.2ZKN December 2009 18 3.9 Observation well 18.4 6.6 3,999

ZF December 2009 18 4.2 Agricultural irrigation 16.1 7.4 1,502

ZG9 December 2009 9 3.7 Observation well 18.5 7.4 1,219

Z06 December 2009 15 2.8 Agricultural irrigation 16.8 7.3 1,378

Z16 December 2009 30 3.2 Agricultural irrigation 17.4 7.7 787


taken using precleaned, brown glass bottles, trans-ported under anaerobic conditions in refrigerated box-es to the laboratory, and stored in the dark under waterat 4 °C until the time of analysis. Water chemistrycharacteristics of the samples are given in Table 2,including temperature (°C), pH, turbidity, specificelectrical conductivity (EC), redox potential (Eh),and dissolved oxygen (DO). These parameters andwater table depth were measured in the field.

In each instance, duplicate water samples for volatileorganic compound (VOC) measurement were collectedand placed in precleaned 40 mL amber glass bottles withTeflon-lined rubber septa. Samples for semivolatile or-ganic compound (SVOC) and OCP measurement werecollected in 1-L amber glass bottles. The bottles werecarefully filled to overflowing, without passing air bub-bles through the sample or trapping air in the sealedbottles. Preparation of bottles included washing withdetergent, rinsing with tap water, ultrapure water(Millipore: Milli-Ro 5 plus and Milli Q plus 185), andacetone (Mallinckrodt Chemical Works St. Louis), andplacing in an oven at 150 °C for 2 h. At each site, HCL (4drops 6N/40mL) was added to thewater sample in orderto bring the solution’s pH down to 2 and prevent bio-degradation and dehydrohalogenation (APHA 1992).

Analytical procedure

The analysis of organic pollutants in the sedimentsamples was performed at the National ResearchCenter for Geoanalysis. PAHs (US EPA Method8310) were tested by high performance liquid chro-matography–mass spectrometry (HPLC-MS), OCPs(US EPA Method 8081A) were measured by gas chro-matography with electron capture detector (ECD), andVOCs (US EPA Method 8260B) were analyzed by

purge and trap extraction systems followed by gaschromatography/mass spectrometry (P&T-GC-MS).Only SVOCs could be detected in the sediment sam-ples and were quantified by HPLC equipped with avariable wavelength fluorescence detector and aSupelcosil LC-PAH (250×4.6 mm i.d., 5 μm particlesize, Supelco) column. The injection volume was5.0 μL, and the column temperature was 30 °C. Thegradient elution program consisted of 65 % water and35 % acetonitrile for 2 min, then 100 % acetonitrile for12 min at a flow rate of 2.0 mL/min.

All organic pollutants in the water samples, includ-ing VOCs, SVOCs, and OCPs, were analyzed in theMinistry of Land and Resources P.R.C. HuadongMineral Resources Supervision and Testing Center(Research Center of Nanjing Institute of Geologyand Mineral Resources).

VOCs in water samples were analyzed by P&T-GC/MS, derived from US EPA method 524.2 (Eichelbergand Bundle 1989). The Tekmar 3000 Purge and Trapautosampler device, operated with Helium as a carrier(gas flux, 50 mL/min; purge time, 11 min), wasconnected to a GC-MS system (HP 6980). ACarbopack C and B (Supelco) trap was used at adesorption temperature of 225 °C and a desorptiontime of 4 min. An HP 5.5 % phenyl methyl siloxanGC column was used for the separation of the targetcompounds (film thickness, 0.25 μm; interior diame-ter, 0.25 mm; length, 60 m). The mass spectrometerwas operated at 315 °C in the selected ion mode. Astock solution of 2,000 μg/mL (EPA 524, Supelco) ofboth fluorobenzene and 1,2-dichlorobenzene-d4 inmethanol was diluted to 200 μg/mL and used as aninternal standard for calibration.

SVOCs (mainly PAHs) in water samples were de-termined by HPLC-MS, derived from US EPA method

Table 2 (continued)

Locationsite

Samplingtime

Well depth(m)

Water depth(m)

Utilization T (°C) pH Turbidity EC(μS/cm)

Eh (mV) DO(mg/L)



ZB December 2009 50 4.0 Domestic water 10.1 7.6 1,164 50.7

DW09 December 2009 16 3.2 Domestic water 15.8 7.0 1,726

DW25 December 2009 30 4.0 Domestic water 7.6 7.6 1,412

SUJL December 2009 SU River water 6.9 7.9 1,085

SU surface water, EC specific electrical conductivity, Eh redox potential, DO dissolved oxygen


610. PAHs in the water samples were analyzed usingan HPLC (Waters 5890) with UV detector and aWaters 3.9×300 mm μBondapak C18 reverse phasecolumn. The HPLC was operated under the followingconditions: a flow rate of 1.8 mL/min, an injectionvolume of 15 μL, a wavelength of 254 nm, and amobile phase of acetonitrile to water of 80:20, andwith isocratic flow conditions. The concentrations of11 PAHs were quantified in this study. According totheir elution orders, they were acenaphthene (Acp),fluorene (Flu), phenanthrene (Phe), anthracene (Ant),fluoranthene (FLT), pyrene (Pyr), benzo(a)anthracene(BaA), chrysene (CHR), benzo(b)fluoranthene (BbF),and benzo(k)fluoranthene (BkF).The detection limitsfor these congeners are 0.2, 0.01, 0.005, 0.01, 0.01,0.005, 0.002, 0.001, 0.001, 0.002, and 0.001 μg/L,respectively.

OCPs and other pesticides in water samples weredetected by GC using a Hewlett Packard GasChromatograph 5890 Series II, supported by a 63NiECD, derived from US EPA Method 8081A. A 30 m×0.53 mm i.d.×0.5 μm film thickness fused silica cap-illary column HP-608 was used for the chromato-graphic separation of pesticides. Helium was used asthe carrier gas and nitrogen as the makeup gas, and theinjection technique was split/splitless. The detectionlimit for OCPs is 0.01 μg/L.

Detailed procedures for sample collection, transpor-tation, extraction, and cleanup were referenced fromthe Geological Survey Standard of GroundwaterPollution (China Geological Survey 2008). Qualitycontrol samples were prepared and analyzed for eachbatch of samples. The QC results showed that thedeviation between duplicates was within 20 % andthe recovery of laboratory control standards wasbetween 80 and 120 %.

Results and Discussion

Concentration variation of organic pollutantsin the core profile

Sixteen SVOCs (Table 1) have been detected in thesediment samples, including phenol-d5, (phenol, 2-fluoro), phenol, (phenol, 3-methyl), nitrobenzene-d5,naphthalene, (naphthalene, 1-methyl), (1,1-biphenylss), dibenzofuran, fluorene, phenanthrene, anthacene,dibutyl phthalate, p-terphenyl-d14, (1,2-benzenedi

acid, disooctyl ester), and benzo(b)fluoranthene. Theconcentrations of SVOCs in the sediments varied to agreat extent at different sampling depths. The totalconcentration of ∑SVOCs in sediments sampled inMay 2009 ranged between 3.8 and 10.4 ng/mg. Asthe bottom of the waste disposal site is at 9.3 m depth,the detected types of SVOCs and their total concen-trations are greater at this depth than at others. Thecore was taken from the base of the landfill (starting at9.3 m depth) for 2.2 m into silty clay (Fig. 3).Therefore, as a whole, the concentrations of differentSVOCs are characterized by a decrease with depth. Itcan be seen from Table 1 that phenol-d5, phenol,2-fluoro, nitrobenzene-d5, 1,1-biphenyl ss, and p-ter-phenyl-d14 were continually detected at all sevendepths. Figure 2 shows the vertical variation of thesefive compounds in the core profile. The total concen-trations of the five compounds are highest (≥9.6 μg/kg)at 10.5–10.7 m depth, and lowest (≥ 3.6 μg/kg) at10.9–11.1 m depth. The silty clay, 3.2 m thick, existsfrom 9.4 to 12.6 m depth and can be regarded as anatural barrier for preventing direct pollution from thelandfill into underlying material.

There are seven PAHs in the detected 16 SVOCs,namely, naphthalene, (naphthalene, 1-methyl), diben-zofuran, fluorene, phenanthrene, anthacene, and benzo(b)fluoranthene. The total PAH concentrations(Table 1) at the different depths range from 0 to990 μg/kg. According to the classification standardsof Maliszewska-Kordybach (1996), sediments withPAH concentrations close to 1,000 μg/kg at 9.3 mdepth, namely, at the bottom of the landfill, are heavilycontaminated. Conversely, the detected PAH concen-trations below 100 μg/kg (∑PAHs) at the remainingsix depths could be indicative of low pollution insediment.

Surface and groundwater

General characteristics

The physicochemical characteristics of the water sam-ples can be seen in Table 2. EC values range from 0.4to 4 mS/cm in groundwater samples, with total dis-solved solid (TDS) between 0.4 and 3.9 g/L, pH 6.6and 8.9, DO 0.5 and8.7 mg/L, and turbidity 0.1 and17.2. EC values range from 0.6 to 3.3 mS/cm insurface water samples, with TDS between 0.6 and1.1 g/L, pH 7.7 and 8.1, DO 3.1 and 5.8 mg/L, and


turbidity up to 90. The groundwater sample (ZKN)in the north of the landfill shows a high mineralcontent (EC ≥4 mS/cm), which is due to a largenumber of anthropogenic sources. The pHs of thegroundwaters at the Zhoukou landfills were slightlyalkaline (Table 2). The mean pHs of groundwaterswere 7.57 and 7.40 in May 2009 and December2009, respectively.

The pulse of oxygen introduced by lowering the watertable likely causes a partial and temporal oxidation ofpreviously reduced species. The dissolved oxygen (DO)values of the wells with 9 m depth are >1 mg/L, indicat-ing the condition is aerobic. The phenolic compoundsgenerally degrade readily under aerobic conditions,while nitrification of ammonium can also occur if oxy-gen is present. However, as nitrifying bacteria growslowly relative to heterotrophic bacteria responsible fordegradation of organic compounds, available oxygenmay be utilized in the degradation of organic substancesthereby preventing nitrification (Keener and Arp 1994).Nitrification has also been observed to be inhibited in thepresence of phenols due to their toxicity (Stafford 1974;Dyreborg and Arvin 1995).

In the interior of the Zhoukou landfill, levels ofalkalinity were very high (average 1,025 mg/L asCaCO3), and they decreased along flow path to about155 mg/L at ZKW. Excess of the alkalinity relative tocalcium (Ca) is likely to be derived from the biodeg-radation of organic matter (Borden et al. 1995;Basberg et al. 1998; Lee et al. 2001). The alkalinityvalues of groundwater samples in the Zhoukou landfillranged between 155 and 2,045 mg/L in May 2009. Asexpected, the nearest well (ZG11) to the landfill showedthe highest values of alkalinity (2,045 mg/L), whichindicates that groundwater near landfill site is beingsignificantly affected by leachate percolation.

Thirty-one organic compounds (out of 92 analyzed)exceeded detection limits. Table 3 shows the detection

rate of the CAHs, MAHs, HAHs, OCPs and otherpesticides, and PAHs in water samples. The detectedresults of these organic compounds in the water sam-ples are shown in Tables 4 and 5. The relative percent-age of the total concentration of the detected pollutantscan be seen from the pie map in Fig. 4, showing PAHsare the main organic contaminants in shallow aquifersaround the Zhoukou landfill.

CAH, MAH, and HAH in water samples

Seven CAHs (out of 29 analyzed) were detected inthe water samples, namely, dichloromethane, chlo-roform, 1,2-dichloroethane, 1,2-dichloropropane,cis-1,2-dichloroethylene, tetrachloroethylene, and1,1,2,2-tetrachloroethane. Besides groundwatersamples ZKN, ZG7, and ZG11 (all ≤0.2 μg/L) fromMay 2009, CAHs were mainly detected in surfacewater samples with concentrations ranging between0.2 and 2.8 μg/L. Comparing these concentrationswith the drinking water quality standards (GB5749-2006) (e.g., chloroform, 60 μg/L; 1,2-dichloropro-pane, 5 μg/L; tetrachloroethylene, 5 μg/L), theCAH concentrations in these water samples arenot above the standards, indicating low levels ofcontamination, probably by industrial discharge insewage water.

Nine MAHs (out of 14 analyzed) were detected,including benzene, toluene, ethylbenzen, ortho-xylene, m+p-xylenes, 1,2,4-trimethylbenzene, isobu-tylbenzene, s tyrene, and isopropylbenzene.Concentrations of the total MAHs ranged from 0.57to 8.96 μg/L in surface water (highest in sewage waterZC), and 0.12 μg/L (Z34, depth 28 m) to 1.14 μg/L(ZG7, depth 9 m) in groundwater samples (mainlyfrom May 2009). Only one HAH (out of 9 analyzed),namely 2-chlorotoluene, was detected in the watersample ZG12.

Table 3 Frequency of detection(%) in water samples

GW groundwater samples, SUsurface water samples

Sampling time Number ofsamples

CAHs MAHs HAHs OCPs+otherpesticides

PAHs

December 2008 10 (GW) 0 20.0 0 0 40.0

3 (SU) 66.7 66.7 0 66.7 66.7

May 2009 17 (GW) 23.5 58.8 5.9 5.9 76.5

6 (SU) 83.3 50.0 0 0 16.7

December 2009 13 (GW) 0 0 0 0 76.9

1 (SU) 0 0 0 100


Tab

le4

CAH,MAH,andHAH

concentrations

inwater

samples

Location

site

Sam

pling

time

CAHs(μg/L)

MAHs(μg/L)

HAHs(μg/L)

DCM

(0.2)a

TCM

(0.1)a

1,2-DCA

(0.2)a

1,2-

DCP

(0.2)a

cis-1,2-

DCE

(0.1)a

PCE

(0.1)a

1,1,2,2-

PCA

(0.1)a

∑CAHs

B(0.2)a

T(0.1)a

E(0.1)a

o-X

(0.1)a

m+p-X

(0.2)a

1,2,4-

TMB

(0.1)a

IBB

(0.1)a

Styrene

(0.1)a

IPB

(0.1)a

∑MAHs

2-Chlorotoluene

(0.1)a

ZA

Decem

ber

2008

--

--

--

--

--

--

--

--

--

-

ZB

Decem

ber

2008

--

--

--

--

--

--

--

--

--

-

ZC

Decem

ber

2008

0.41

0.63

0.66

0.47

-0.63

-2.8

-8.15

0.17

0.22

0.42

--

--

8.96

-

ZD

Decem

ber

2008

--

--

--

--

--

--

--

--

--

-

ZE

Decem

ber

2008

--

1.3

--

0.25

-1.55

-1.33

0.16

--

--

--

1.49

-

ZF

Decem

ber

2008

--

--

--

--

--

--

--

--

--

-

ZG

Decem

ber

2008

--

--

--

--

--

--

--

--

--

-

Z6A

Decem

ber

2008

--

--

--

--

--

--

--

--

--

-

Z6

Decem

ber

2008

--

--

--

--

--

--

--

--

--

-

SW29

Decem

ber

2008

--

--

--

--

-0.51

--

--

--

-0.51

-

SW49

Decem

ber

2008

--

--

--

--

--

--

--

--

--

-

SW59

Decem

ber

2008

--

--

--

--

--

--

--

--

--

-

ZK1

Decem

ber

2008

--

--

--

--

-0.44

--

--

--

-0.44

-

ZKE

May

2009

--

--

--

--

--

--

--

--

--

-

ZKW

May

2009

--

--

--

--

-0.13

--

--

--

-0.13

-

ZKS

May

2009

--

--

--

--

--

--

--

--

--

-

ZKN

May

2009

--

--

0.15

--

0.15

0.44

--

--

--

-0.13

0.57

-

ZF

May

2009

--

--

--

--

--

--

--

--

--

-

ZG7

May

2009

--

--

0.19

--

0.19

0.97

--

0.17

--

--

-1.14

-

ZG9

May

2009

--

--

--

--

-0.19

--

--

--

-0.19

-

ZG10

May

2009

--

--

--

--

--

--

--

--

--

-

ZG11

May

2009

--

--

--

0.16

0.16

0.55

--

--

--

--

0.55

-

ZG12

May

2009

--

--

--

--

0.37

--

--

--

--

0.37

0.26

Z16

May

2009

--

--

--

--

-0.24

--

--

--

-0.24

-

Z34

May

2009

--

--

--

--

-0.12

--

--

--

-0.12

-


Tab

le4

(con

tinued)

Location

site

Sam

pling

time

CAHs(μg/L)

MAHs(μg/L)

HAHs(μg/L)

DCM

(0.2)a

TCM

(0.1)a

1,2-DCA

(0.2)a

1,2-

DCP

(0.2)a

cis-1,2-

DCE

(0.1)a

PCE

(0.1)a

1,1,2,2-

PCA

(0.1)a

∑CAHs

B(0.2)a

T(0.1)a

E(0.1)a

o-X

(0.1)a

m+p-X

(0.2)a

1,2,4-

TMB

(0.1)a

IBB

(0.1)a

Styrene

(0.1)a

IPB

(0.1)a

∑MAHs

2-Chlorotoluene

(0.1)a

Z40

May

2009

--

--

--

--

-0.2

--

--

--

-0.2

-

DW09

May

2009

--

--

--

--

--

--

--

--

--

-

DW18

May

2009

--

0.97

1.25

--

-2.22

0.5

--

--

--

--

0.5

-

DW23

May

2009

--

--

--

--

--

--

--

--

--

-

DW25

May

2009

--

--

--

--

--

--

--

--

--

-

ZC

May

2009

-0.26

0.38

0.38

-0.61

-1.63

-5.43

1.14

-1.45

0.13

0.13

0.59

-8.87

-

ZD

May

2009

--

--

-0.22

-0.22

--

--

--

--

--

-

SUJL

May

2009

-1.33

--

--

-1.33

--

--

--

--

--

-

SULG

May

2009

--

--

--

--

-1.07

--

--

--

-1.07

-

SUY

May

2009

-1.81

--

--

-1.81

--

--

--

--

--

-

SULD

May

2009

--

1.11

--

--

1.11

-0.57

--

--

--

-0.57

-

NoCAH,MAH

andHAH

hasbeen

detected

inwater

samples

collected

inDecem

ber20

09

CAHschlorinatedaliphatic

hydrocarbo

ns(including

DCM

dichloromethane,TCM

chloroform

,1,2-DCA1,2-dichloroethane,1,2-DCP1,2-dichloroprop

ane,

cis-1,2-DCEcis-1,2-

dichloroethy

lene,PCEtetrachloroethylene,

1,1,2,2-PCA1,1,2,2-tetrachloroethane),MAHsmon

ocylic

arom

atic

hydrocarbo

ns(including

Bbenzene,

Ttoluene,

Eethy

lbenzen,

o-X

ortho-xy

lene;m+p-Xm+p-xy

lenes;1,2,4-TMB1,2,4-trim

ethy

lbenzene,IBBisob

utylbenzene,IPBisop

ropy

lbenzene),HAHshalogenatedarom

aticshy

drocarbo

nsaDetectio

nlim

it


Tab

le5

OCPandPA

Hconcentrations

inwater

samples

Location

site

Sam

pling

time

OCPs(μg/L)

PAHs(μg/L)

BHC

(0.01)

aγ-chlordane

(0.01)

aEndosulfan-I

(0.01)

aEndosulfan

sulfate

(0.01)

a

Nap

(0.2)a

Acp

(0.01)

aFlu

(0.005)a

Phe

(0.01)

aAnt

(0.01)

aFLT

(0.005)a

Pyr

(0.002)a

BaA

(0.001)a

CHR

(0.001)a

BbF

(0.002)a

BkF

(0.001)a

∑PA

Hs

ZA

Decem

ber

2008

--

--

-0.38

0.682

0.519

0.168

0.261

0.158

0.0181

--

-2.19

ZB

Decem

ber

2008

--

--

--

--

--

--

--

--

ZC

Decem

ber

2008

-0.058

--

--

0.0477

0.125

-0.0323

0.0306

--

--

0.24

ZD

Decem

ber

2008

--

--

--

--

--

--

--

-

ZE

Decem

ber

2008

-0.014

--

--

0.0184

0.043

-0.0097

0.007

--

--

0.08

ZF

Decem

ber

2008

--

--

-0.215

0.362

0.452

0.079

0.153

0.0815

0.0086

--

-1.35

ZG

Decem

ber

2008

--

--

--

0.0241

0.033

--

--

--

-0.06

Z6A

Decem

ber

2008

--

--

--

--

--

--

--

--

Z6

Decem

ber

2008

--

--

--

--

--

--

--

--

SW29

Decem

ber

2008

--

--

--

--

--

--

--

--

SW49

Decem

ber

2008

--

--

0.6

--

--

--

--

--

0.60

SW59

Decem

ber

2008

--

--

--

--

--

--

--

--

ZK1

Decem

ber

2008

--

--

--

--

--

--

--

--

ZKE

May

2009

0.18

--

--

0.147

0.3642

0.8299

0.174

0.2984

0.1763

0.0334

0.0396

0.0131

0.0042

2.08

ZKW

May

2009

--

--

--

0.2793

0.645

0.122

0.3549

0.1691

0.029

0.0363

--

1.64

ZKS

May

2009

--

--

-0.144

0.2971

0.5953

0.117

0.2096

0.1224

0.0139

--

-1.50

ZKN

May

2009

--

--

-0.06

0.177

0.2833

0.038

0.0531

0.0295

--

--

0.64

ZF

May

2009

--

--

-0.212

0.4391

0.7569

0.168

0.2753

0.1578

0.0214

0.0207

--

2.05

ZG7

May

2009

--

--

--

0.0259

0.028

--

--

--

-0.05

ZG9

May

2009

--

--

--

0.02

0.024

--

--

--

-0.04

ZG10

May

2009

--

--

--

0.0338

0.028

--

--

--

-0.06

ZG11

May

2009

--

--

-0.02

0.0395

0.0312

--

--

--

-0.09

ZG12

May

2009

--

--

--

0.0215

0.0189

--

--

--

-0.04

Z16

May

2009

--

--

-0.115

0.2596

0.3552

0.075

0.1474

0.097

0.0093

0.0099

--

1.07

Z34

May

2009

--

--

-0.129

0.2818

0.452

0.104

-0.0943

0.0118

0.0098

--

1.08


Tab

le5

(con

tinued)

Location

site

Sam

pling

time

OCPs(μg/L)

PAHs(μg/L)

BHC

(0.01)

aγ-chlordane

(0.01)

aEndosulfan-I

(0.01)

aEndosulfan

sulfate

(0.01)

a

Nap

(0.2)a

Acp

(0.01)

aFlu

(0.005)a

Phe

(0.01)

aAnt

(0.01)

aFLT

(0.005)a

Pyr

(0.002)a

BaA

(0.001)a

CHR

(0.001)a

BbF

(0.002)a

BkF

(0.001)a

∑PA

Hs

Z40

May

2009

--

--

--

0.0137

--

0.0818

0.1471

0.0095

--

-0.25

DW09

May

2009

--

--

--

--

--

--

--

--

DW18

May

2009

--

--

--

--

--

--

--

--

DW23

May

2009

--

--

--

--

--

--

--

--

DW25

May

2009

--

--

--

--

--

--

--

--

ZC

May

2009

--

--

--

0.0465

0.1189

-0.0581

0.0348

0.0077

0.0134

--

0.28

ZD

May

2009

--

--

--

--

--

--

--

--

SUJL

May

2009

--

--

--

--

--

--

--

--

SULG

May

2009

--

-0.088

--

--

--

--

--

--

SUY

May

2009

--

--

--

--

--

--

--

--

SULD

May

2009

--

0.021

--

--

--

--

--

--

-

ZKE

Decem

ber2009

--

--

--

0.0322

0.017

-0.016

0.0097

--

--

0.07

ZKW

Decem

ber2009

--

--

--

0.0311

0.066

0.015

0.0327

0.0213

--

--

0.17

ZKS

Decem

ber2009

--

--

-0.025

0.0869

0.103

0.041

0.1135

0.057

0.0077

0.0076

--

0.44

ZKN

Decem

ber2009

--

--

-0.014

0.0491

0.108

0.023

0.0339

0.026

--

--

0.25

ZF

Decem

ber2009

--

--

-0.013

0.0506

0.142

0.027

0.0272

0.0491

0.0344

0.0053

--

0.35

ZG9

Decem

ber2009

--

--

--

0.0916

0.196

0.028

0.0301

0.0195

--

--

0.37

Z06

Decem

ber2009

--

--

--

--

--

--

--

--

Z16

Decem

ber2009

--

--

--

0.0625

0.156

0.026

0.0457

0.0361

0.0033

0.0033

--

0.33

Z34

Decem

ber2009

--

--

--

--

-0.0323

0.0466

--

--

0.08

Z40

Decem

ber2009

--

--

--

0.0732

0.171

0.033

0.0661

0.0525

0.005

0.0058

--

0.41

ZB

Decem

ber2009

--

--

--

--

--

0.006

--

--

0.01

DW09

Decem

ber2009

--

--

--

--

--

--

--

--

DW25

Decem

ber2009

--

--

--

--

--

--

--

--

SUJL

Decem

ber2009

--

--

--

--

--

0.008

--

--

0.01

aDetectio

nlim

it

OCPsorgano

chlorine

pestcides(BHC

benzenehexachloride),PA

Hspo

lycyclic

arom

atic

hydrocarbo

ns(Nap

naph

thalene,

Acp

acenaphthene,Flu

fluo

rine,Phe

phenanthrene,Ant

anthracene,FLT

fluo

ranthene,Pyr

pyrene,BaA

benzo(a)anthracene,CHRchrysene,BbF

benzo(b)fluo

ranthene,BkF

benzo(k)fluo

ranthene)


BTEX compounds (benzene, ethylbenzene, tolu-ene, and three isomers of xylene) of MAHs are clas-sified as environmental priority pollutants, which maynot exceed 10, 700, 300, and 500 μg/L in drinkingwater, respectively, according to the National Chinese(NC) standards (GB-5749-2006). They are commonlyfound together in crude petroleum and petroleumproducts such as gasoline and diesel fuel. The pres-ence of these hydrocarbons in the environment is ahazard to public health and an ecological concern, dueto their toxicity and ability to bioaccumulate throughthe food chain (Brigmon et al. 2002). BETX areprominent components of gasoline, and their presencein water is usually an indication of gasoline contami-nation. The contaminants in this study do not exceedthe defined limits for Chinese drinking water stand-ards, which is the same with the World HealthOrganization (WHO) guidelines for drinking waterquality (WHO 2006).Toluene is mainly detected inthe water samples from the Low-Lying Gully andsewage ditch (ZC), xylene detected in the sewageditch (ZC), and benzene mainly detected in ground-water samples (such as ZG7, ZKN, ZG11, and ZG12at 9 m depth) close to the landfill.

The detection rate of BTEX in May 2009 is higherthan that in December 2009. BTEX concentrations atindividual boreholes were highly variable over time.These temporal variations appeared to result from notonly hydraulic variations and seasonal groundwaterflow variations but also preferential dissolution andbiodegradation. In particular, systematic decreases inSO4

2− concentrations, increases in HCO3− concentra-

tions, and the presence of degradation products inregions of the plume where BTEX concentrations arerelatively high are indicative of degradation processes(Grbic-Galic and Vogel 1987; Wiedemeier et al.1995). Ranking the hydrocarbon compounds by theplume-scale degradation rate estimate, from highest tolowest rate gave the order: toluene, o-xylene, naphtha-lene, m- and p-xylene, trimethylbenzene, ethylben-zene, and benzene (Davis et al. 1999). For anaerobic, nitrate-rich BTEX contaminated aquifer,Daniel and Borden (1997) found highest degradationrates near the source of their plume, with decreasingdegradation rates with distance down the plume.During biodegradation, microorganisms transformavailable carbon into forms useful for energy and cellproduction. This results in oxidation of the electron

Fig. 4 Total frequency ofthe detected organic com-pounds in the groundwatersamples, like CAHs, OCPs,MAHs, and PAHs


donor (such as organic matter) and reduction of elec-tron acceptor [such as DO, nitrate, iron (III) or Mn(III), sulfate, and carbon oxide] (Essaid et al. 1995; Luet al. 1999). The aquifer near the landfill is character-ized by elevated BTEX concentrations in May 2009,relatively low concentrations of nitrate, and sulfate. Incontrast, the distribution area of low BTEX concen-trations (Fig. 5e) under aerobic condition (DOdetected over 1 mg/L) and nitrate (Fig. 5b), sulfate(Fig. 5c), and manganese (Fig. 5d) concentrations ishigh. The in situ microbes seem to be using fuelhydrocarbons as their carbon and energy sources,thereby contributing to the natural removal process.

OCPs and other pesticides in water samples

From Table 5, it can be seen that only one OCP (out of11 analyzed), namely, benzene hexachloride wasdetected in the groundwater sample ZKE (27 m depth)with a concentration of 0.18 μg/L, which is far belowthe Chinese drinking water standard of 5 μg/L. Withthe exception of well ZKE, groundwater samples arenot contaminated by OCPs and other pesticides, whichhave been detected in drainage canals in the east of thelandfill and in water from the LuDong Trunk Canaland the Low-Lying Gully. This indicates that, except

for local point polluted by OCPs in the eastern part ofthe landfill, there is less pesticide contribution to theshallow aquifer. Three kinds of common pesticides(out of 13 analyzed), including γ-chlordane (detectedin ZC and ZE), endosulfan-I and endosulfan sulfate,were detected in water sample from the sewage ditchwith concentrations varying from 0.021 to 0.088 μg/L,indicating that local agricultural activities contributeirrigation return flows to the sewage water. This is alsoa potential source for some shallow groundwater pol-lution due to the untreated bed of the sewage ditch.The sources of pesticide residues in the waters studiedare agricultural practices within the study area, incombination with rainfall. Maximum concentrationsof this compound were detected in May, possiblydue to surface run-off. The spatial and temporal dis-tribution of pesticides obtained from the monitoringnetwork shows no clear trends for prediction of futureconcentrations. Nitrate–N concentrations and pesticidedetections show no clear relationship, suggesting dif-ferent source, transport, or degradation pathways.

PAHs distribution in water samples

PAHs are the main contaminants of concern detectedin the water samples. Unlike CAHs, MAHs, HAHs,

Fig. 5 Concentration contours for the chloride (a), nitrate (b), sulfate (c), manganese (d), BTEX (e), and PAHs (f) of groundwatersamples in May 2009. Units are micrograms per liter for BTEX and PAHs and milligrams per liter for the rests


and OCPs, most groundwater samples containedPAHs in the summer and winter 2009 samples, witha high detection rate of more than 75 % (Table 3).PAH concentrations in nine representative wells(Fig. 6) varied from 0.04 to 2.08 μg/L in May 2009and from 0.07 to 0.44 μg/L in December 2009, indi-cating that rainfall can enhance contaminant leakagefrom the waste disposal site into groundwater duringthe summer. It is clear that PAH concentrations insummer are higher than those in winter. Many of thePAH compounds in water samples were present atconcentrations in excess of 1 μg/L, suggesting thatwater in this area was heavily contaminated (Zhouand Maskaoui 2003), especially in May 2009. In thisstudy, the heavily contaminated water samples, suchas ZA and ZF sampled in December 2008, and ZKE,ZKS, ZKW, ZF, Z16, and Z34 sampled in May 2009,mainly occur at water depths of 18–30 m, in somecases with concentrations of total PAHs (in ZKE andZF) beyond 2 μg/L (Fig. 7). The detectable propor-tions of three- and four-ring PAHs were the highest,with two- and five-ring PAHs the lowest (Fig. 6).Univariate Pearson correlation matrix (Table 6) showsa good correlation among all PAHs except BaA.Higher Pearson coefficients of different PAHs underthe 0.01 significant level can reflect the similar pollu-tion sources, e.g., Acp, Ant, FLT, and Pyr.

Eleven PAHs (out of 16 analyzed) were detected,including Nap, Acp, Flu, Phe, Ant, FLT, Pyr, BaA,CHR, BbF, and BkF. No six-ring PAHs were detectedin water samples. The detected concentrations in water

is 0.6 μg/L for two-ring PAHs (Nap), ranged 0.01–1.75 μg/L for three-ring PAHs (Acp, Flu, Phe, andAnt), 0.01–0.59 μg/L for four-ring PAHs (FLT, Pyr,BaA, and CHR), and 0.02 μg/L for five-ring PAHs(BbF and BkF). The total concentrations of these 11PAHs in water ranged from 0.01 μg/L at well ZB to2.19 μg/L at well ZA, with a mean concentration of0.62 μg/L (Table 5). The three dominant PAHs foundin most groundwater samples are Phe, Flu, and FLTin the study area (Fig. 8). They formed 0–52(mean, 35)%, 0–46 (mean, 19)%, and 0–40 (mean,17)% of the total PAHs, respectively. The lowermolecular weight (LMW, two to three rings) PAHsdominate in all samples with the exception of Z40.Generally, PAHs from a petrogenic source show adepletion of higher molecular weight (HMW, fourto six rings) PAHs relative to LMW PAHs, whilepyrogenic sources are abundant in HMW PAHs(Zakaria et al. 2002). Most groundwater aroundthe Zhoukou landfill not only contains a consider-able amount of LMW PAHs but is also abundant inHMWPAHs, indicating the input of both petrogenic andpyrogenic origins.

Source and degradation of PAHs

Source diagnosis by diagnostic ratios of PAHs

Inferring the sources of PAHs is widely considered tobe very important to study the transport and fate ofPAHs in environment (Wan et al. 2006). Generally,ratios of various PAH concentrations have usuallybeen undertaken to diagnose the possible sources ofPAHs (Fernandes et al. 1997; Yunker et al. 2002).When the concentrations of different PAHs in water

Fig. 6 Non-outlier range, interquartile range (IQR), and medianconcentrations (box-and-whisker plots) of ∑11PAHs in shallowgroundwater samples (ZKE, ZKS, ZKW, ZKN, Z16, Z34, Z40,ZF and ZG9) during the two sampling campaigns

Fig. 7 Plots (a) of PAHs concentrations corresponding togroundwater sampling depth


samples reaches the quantitation limit, some selecteddiagnostic ratios have been calculated and shown inTable 7. Most of the water samples detected PAHs inthis study show these ratios intermediate between

pyrogenic and petrogenic values, compared with thereported values for particular processes (Table 8).

Petroleum often contains more thermodynamicallystable compounds such as Nap, Flu, Phe, and CHR,

Table 6 Pearson coefficient of different PAHs in the groundwater samples (n026) around the Zhoukou landfill

Acp Flu Phe Ant FLT Pyr BaA CHR PAHs

Acp 1

Flu 0.99a 1

Phe 0.69b 0.86a 1

Ant 0.80a 0.90a 0.95a 1

FLT 0.76b 0.80a 0.92a 0.93a 1

Pyr 0.75a 0.71a 0.94a 0.96a 0.91a 1

BaA -0.10 0.28 0.51 0.47 0.51 0.51 1

CHR 0.60 0.71b 0.88a 0.84a 0.94a 0.94a 0.67b 1

PAHs 0.88a 0.95a 0.97a 0.98a 0.93a 0.84a 0.47 0.86a 1

a Correlation is significant at the 0.01 level (two-tailed)b Correlation is significant at the 0.05 level (two-tailed)

Fig. 8 Distribution of organic compound groups identified inwater samples around the landfill a in May 2009 and b inDecember 2009. Results are based on GC peak areas of the

GC-MS full scan analysis. ∑PAHs denotes total PAH concen-tration, i.e., the sum of the individual mass concentrations of the11 PAH congeners


while FLT and Pyr are usually the most abundantcompounds for pyrolytic PAHs (Doong and Lin2004). The FLT/Pyr and FLT/(FLT + Pyr) ratios can beuseful tools to check PAHs pollution origin (Gschwendand Hites 1981; Gogou et al. 1998; Magi et al. 2002).Values of FLT/Pyr <1, FLT/(FLT + Pyr) <0.5, Phe/Ant >15, and CHR/BaA >1 indicate a petrogenicorigin of contamination (De Luca et al. 2004). Inthis context, Z40 (May), Z34 (December), ZF(December) show a strong petrogenic characterand that FLT/Pyr and FLT/(FLT + Pyr) ratios aremuch below 1 and 0.5, respectively, where therests show ratios compatible with pyrogenic sour-ces of contamination, probably originate mainlyfrom grass, wood, and coal combustion.

Similar results were observed for the CHR/BaAratios. CHR and BaA are both derived from the com-bustion processes with CHR/Pyr ratio lower than 1.This ratio in this study ranges between 0.15 and 1.74,indicating combustion processes and petroleum hydro-carbons are the possible main source of PAHs in watersamples in this study. The ratio Ant/(Ant + Phe) has

been suggested as diagnostic indicator for distin-guishing between pyrogenic and petrogenic sourceswith values >0.1 indicating pyrolytic souces, where-as <0.1 suggest petrogenic (Budziński et al. 1997).In this study, this ratio is >0.1 for the water samples(e.g., ZA, ZF, ZKE, and ZKS in Table 6) where Antand Phe are detectable, indicating dominance of fuelcombustion and coal burning processes in thesesample sites. Most samples with Phe/Ant <10 andFLT/Pyr <1 were characterized as a mixture of py-rolytic and petrogenic contamination, which is ingood agreement.

The distribution of LMWand HMW PAHs is also atool for identifying the petrogenic/pyrolytic origin ofPAHs (Sicre et al. 1987; Budziński et al. 1997). Thehigher the LMW/HMW ratio is, the higher the preva-lence of petrogenesis on pyrolytic origin of PAHs is(De Luca et al. 2004). The LMW/HMW ratios of thecollected water samples range from 0.06 to 8.34 withmean value of 3.24. Except for Z34 (May) and Z40(December), Table 3 also shows that LMW are clearlypredominant over HMW, suggesting a definite

Table 7 Diagnostic ratios used with their typically reported values for particular processes

PAH ratio Value range Source Reference This study

∑LMW/∑HMW <1 Pyrogenic Zhang et al. 2008; Budzińskiet al. 1997

0.06–8.34

>1 Petrogenic

Flu/(Flu+Pyr) <0.5 Petrol emissions Ravindra et al. 2008b 0.09–0.86

>0.5 Diesel emissions

Ant/(Ant+Phe) <0.1 Petrogenic Pies et al. 2008 0.12–0.29

>0.1 Pyrogenic

FLT/(FLT+Pyr) <0.4 Petrogenic Gogou et al. 1998; De LaTorre-Roche et al. 2009

0.36–0.68

0.4–0.5 Fossil fuel combustion

>0.5 Grass, wood, coal combustion

FLT/Pyr <1 Petrogenic Sicre et al. 1987; Baumardet al. 1998a; b

0.55–2.1

>1 Pyrogenic

BaA/(BaA+CHR) <0.2 Petrogenic Akyüz and Çabuk 2010;Yunker et al. 2002;Wang et al. 2010

0.37–0.87

0.2–0.35 Coal combustion

>0.35 Combustion

CHR/BaA <1 Pyrogenic Soclo et al. 2000 0.15–1.74

>1 Petrogenic

Phe/Ant <10 Pyrogenic Baumard et al. 1998 a;b; Cao et al. 2005

2.51–7.46

>15 Petrogenic


petrogenic origin of PAHs. The low solubilities ofLMW PAHs compared to the HMW compounds(Lee and Lee 2004; Sahu et al. 2004) may also beresponsible for the high ratios in water, as opposed tosource material alone (Tobiszewski and Namieśnik2012). These results showed that the PAHs contami-nation in this study was probably from mixture sour-ces of petroleum and combustion products.

Source apportion by principal component analysis

To provide insight into the accuracy and quantificationof source apportion, principal component analysis(PCA) was applied to analyze the data set. PCAreduces the number of variables in the original dataset into principal components without significant lossin the total variance of the data. The loading that eachvariable in the original data contributes to the principalcomponents enables grouping of data with similarbehaviors. Values below the detection limit werereplaced by half of the method detection limits forthe statistical analysis. The score and loading plots

(Fig. 9) obtained by PCA can show the similaritiesor dissimilarities between ambient PAH profiles.

After autoscaling, two significant components wereidentified, giving account for 62.5 and 16.3 % of thetotal variance, respectively. The third component takesinto account only 8.9 % of the total variance and wasnot considered in the present analysis. Figure 9ashows the loading plot and substantiates that the firstcomponent is mainly related to Phe, FLT, Ant, Pyr,Acp, Flu, and BaA, whereas the second componentis mainly related to BkF, BbF, and CHR. Highloads of Pyr, Phe, and FLT might be an indicationof diesel combustion (Ravindra et al. 2008a). Thereare also three groups identified on the factor scoreplot (Fig. 9b). Group 1 clusters samples (such asZG7, ZG9, and ZG10) mostly collected around thelandfill in Decemeber; group 2, samples collectedfrom ZKN, ZKW, Z16, and Z34 in May and ZF inDecember; and group 3 only contains one sample,which collected from the ZA. The discrimination inthree groups was confirmed by hierarchical cluster-ing analysis (Fig. 10).

Table 8 PAH characteristics and diagnostic ratios from water samples around the Zhoukou landfill

ID ∑PAHs LMW/HMW

LMW% HMW% Phe/Ant

Flu/(Flu+Pyr)

FLT/Pyr

FLT/(FLT+Pyr)

CHR/BaA

Ant/(Ant+Phe)

BaA/(BaA+CHR)

Water samples in May 2009

ZKE 2.08 2.68 72.8 27.2 4.77 0.67 1.69 0.63 1.19 0.17 0.46

ZKS 1.50 3.33 76.9 23.1 5.09 0.71 1.71 0.63 0.16

ZKW 1.64 1.89 64.0 33.8 5.29 0.62 2.10 0.68 1.25 0.16 0.44

ZKN 0.64 6.76 87.1 12.9 7.46 0.86 1.80 0.64 0.12

ZC 0.28 1.45 59.2 40.8 0.57 1.67 0.63 1.74 0.36

ZF 2.05 3.32 76.8 23.2 4.51 0.74 1.74 0.64 0.97 0.18 0.51

Z16 1.07 3.17 75.3 23.7 4.74 0.73 1.52 0.60 1.06 0.17 0.48

Z34 1.08 8.34 89.3 10.7 4.35 0.75 0.83 0.19 0.55

Z40 0.25 0.06 5.4 94.6 0.09 0.56 0.36

Water samples in December 2009

ZG9 0.37 6.36 86.4 13.6 7.00 0.82 1.54 0.61 0.13

Z34 0.08 0.00 0.0 100.0 0.69 0.41

ZF 0.35 2.01 66.7 33.3 5.26 0.51 0.55 0.36 0.15 0.16 0.87

ZKN 0.25 3.24 76.4 23.6 4.70 0.65 1.30 0.57 0.18

ZKE 0.07 1.91 65.7 34.3 0.77 1.65 0.62

ZKS 0.44 1.38 57.9 42.1 2.51 0.60 1.99 0.67 0.99 0.28 0.50

ZKW 0.17 2.08 67.5 32.5 4.40 0.59 1.54 0.61 0.19

Z40 0.41 2.24 68.2 30.4 5.18 0.58 1.26 0.56 1.16 0.16 0.46

Z16 0.33 2.87 73.4 25.6 6.00 0.63 1.27 0.56 1.00 0.14 0.50


Fig. 9 Principal component analysis (PCA) loadings for PAHsin groundwater samples in the Zhoukou landfill site: a Principalcomponents loading plot and b component score plot. Principal

component 1 and 2 (PC1 and PC2) account for 62.5 and 16.3 %of the variance in the data set, respectively (ZKE not reportedhere because located outside the graph b)

Fig. 10 Hierarchical clus-tering of the PAHs concen-trations from water samplesaround the Zhoukoulandfill


The samples circled in the group 1 contain similarcontaminants, particularly Nap, with a high contribu-tion relative to the other PAHs. The samples in thegroup 1are mostly those collected around the landfill.Samples in group 2 were mostly collected in May2009 and show similar characteristics, which may berelated to the origins of contamination and timing ofthe sampling. These samples are characterized by pos-itive loading in the PC2 and negative loading in PC1,so have more contribution from Phe, Ant, FLT, Pyr,Flu, and Acp the than other samples. Although thesample of ZA of the group 3 was collected from theshallow aquifer west of the landfill with the highesttotal PAHs concentration (up to 2.19 μg/L); CHR,BaA, and Flu make up a relatively high contributionto ZA.

Degradation of PAHs

Although PAH may undergo adsorption, volatiliza-tion, photolysis, and chemical degradation, microbialdegradation is the major degradation process (Zaidiand Imam 1999; Christensen et al. 2001; Haritash andKaushik 2009). Sorption of leachate organic matter onto aquifer material seems to be of only minor signifi-cance according to column experiments reported in theliterature (Christensen 1992; Rügge et al. 1995). Thebiodegradation of PAHs has been observed under bothaerobic and anaerobic conditions (Haritash andKaushik 2009), which depends on the environmentalconditions, number and type of the microorganisms,and nature and chemical structure of the chemicalcompound being degraded. In this study, the DO con-centrations of most groundwater samples in excess of1.0 mg/L are identified that the aquifer is aerobic. Thevalue of 1.0 mg/L is defined in order to minimize thepresence of nitrate-reducing microenvironments in theaerobic aquifer (Lyngkilde and Christensen 1992).The pH values are lower in winter season than thatin summer season. Although altering the pH of waterfrom neutral to pH 6.0 and 8.0 had very little or noeffect on Phe (Zaidi and Imam 1999), the increasingpH values of water in summer season may have effecton reducing the other PAHs degradation.

The monitoring results show the water table is higherin December than that in May, due to the reduction ingroundwater exploitation over the winter season. Littleis known about the effects of fluctuations of the water

table on organic contaminant degradation under fieldsituations. From laboratory experiments, it was demon-strated that fluctuating the water table enhanced thedegradation of diesel oil (Rainwater et al. 1993). It canbe expected that the dynamics regime imposed by afluctuating water table and the resulting differences inunsaturated zone (e.g., soil aeration), not only affect themicrobial and chemical reactions that organic pollutantsundergo but also the transport of gases and solutesthrough the aquifer (Sinke et al. 1998).

Some studies showed that some LWM PAHs arebiosusceptible and can be biodegraded more rapidlythan the HMW PAHs (Hinga 2003; Rothermich etal. 2002). PAHs are known to dissipate undernitrate- and sulfate-reducing conditions; sometimes,HMW PAHs after LMW PAHs have been utilized/degraded (Meuller et al. 1989), while the presenceof phenanthrene is reported to inhibit degradationof pyrene (McNally et al. 1999). Because HMWPAHs (such as Pyr, BaA, and BbF) are more resis-tant to microbial degradation processes, they tendto persist longer in contaminated environments (vanBrummelen et al. 1998; Neilson and Allard 1998;Bosma et al. 2001), and their degradation pathwaysare less well understood.

The Gibbs free energy for oxidation of organiccarbon decreases at neutral pH in the order: O2,NO3

−, MnO2, Fe(OH)3, SO42−, and CO2 (van

Breukelen 2003; Wilson et al. 2004). Therefore, aero-bic degradation followed by nitrate reduction oxidizesorganic carbon at the fringes of plumes. Figure 11shows the chloride levels and the concentrations ofredox-sensitive parameters with BTEX and PAHs con-centrations along the NS and WE cross-sections in theZhoukou landfill. A combination of geochemical data,in terms of changes in solute concentrations along theflowpath, and the PAHs concentrations in each wellcan be used to describe contaminated extent in theshallow aquifer system. As groundwater moves awayfrom the landfill the following changes occur:

1. Chloride is not considered to undergo any chem-ical or physico-chemical reactions in the aquifersand as such is considered inert or conservative(Christensen 1992). For this reason, chloride canbe used to study dispersion and dilution of acontaminant plume. Measured chloride concentra-tion in groundwater samples was 239 mg/L


ranging between 51 and 976 mg/L in May 2009.There was a definite attenuation pattern observedin wells down gradient of the landfill site (Fig. 4a).

2. Sulfate concentrations increase as groundwatermoves outside the sulfate reduction zone. A de-cline of sulfate near the landfill is noted due to

Fig. 11 Chloride and redox sensitive species with main organiccontaminants along the NS and WE cross-sections in the Zhou-kou landfill. 1Middle and fine sand, 2 silt, 3 silty clay, 4 fill soil

(clayey silt), 5 landfill, 6 groundwater table, 7 groundwater flowdirection, 8 monitoring well with well screen


sulfate reduction. Sulfate concentration in landfillleachate (<200 mg/L at the monitoring wells of27 m depth) is generally too low to maintain adegradation potential equal to iron reduction.

3. Groundwater near the landfill contains elevatedconcentrations of dissolved Fe and Mn, mainlymobilized under reducing conditions from landfillleachate.

4. The presence of ammonium (NH4+) in groundwa-

ter has biochemical significance as a useful indi-cator of organic pollution (Chapman 1992).Ammonium appears to give way downstream toa narrow zone in which elevated concentrations ofnitrite are detected (e.g., at Z40, ZKW, and Z16).This may be interpreted as partial nitrification ofthe ammonium plume due to infiltrating oxic wa-ter. Downstream of the nitrite well, nitratebecomes the dominant species, suggesting furthernitrification. Nitrate reduction causes disappear-ance of nitrate with depth upstream from the land-fill (e.g., Postma et al. 1991), such as themonitoring wells at 9 m depth (ZG10, ZG11, andZG12). Nitrate reduction is a likely process at themixing zone of landfill leachate and shallow ni-trate containing groundwater.

Nitrate >1 mg/l (with a maximum of 487 mg/L atDW18) was encountered downstream of the landfill,and nitrate <0.1 mg/L water samples were found nearthe landfill at wells ZG10, ZG11, ZG12, ZKE, ZKN,and ZF, especially ~9 m depth, indicating that denitri-fication is likely a dominant redox process at thedownstream fringes of the plume. The contributionof aerobic and nitrate-reducing zones to natural atten-uation of a plume increases with the O2 and NO3

−

concentrations in pristine groundwater and the extentof mixing with the leachate plume (van Breukelen2003). Figure 4c shows lower NO3

− concentrationsnear the landfill, and there are higher Phe concentra-tions detected in ZKE, ZKW, ZKN, etc (Table 4),which are the monitoring wells near the landfill. Thisis consistent with that the idea that a lack of N mayslow down the biodegradation of phenanthrene (Zaidiand Imam 1999) and cause it to accumulate in thegroundwater.

Most groundwater samples (such as ZKN, ZKE,ZF, and Z16) in December have higher NO3

− concen-trations and lower pH values than in May. Thesewill enhance degradation of PAHs due to the added

inorganic nitrogen under aerobic conditions and resultin lower PAHs concentrations in winter season. Inaddition, the higher HCO3

− concentrations (mean val-ue, 926 mg/L at 27 m depth) in December than in May(mean value, 487 mg/L) support this conclusion, asHCO3

− is probably a by-product of PAHs degradation.

Characteristics of organic pollution in water bodies

Horizontal distribution of organic pollution

No organic compounds were detected in samples fromwells ZB, Z6A, Z6, SW59, DW23, DW09, DW25,and Z06. The contamination plume is identifiable bythe dashed oval-shaped line in Fig. 1 and is distributedacross the area where groundwater depth is 2–4 m. Inthe horizontal direction, the extent of pollution aroundthe landfill is 2 km from south to north and 3 km fromeast to west. In the contaminated area, relatively fewerorganic compounds are detectable in December asopposed to May. Theoretically, VOCs are not oftenfound in surface waters (especially lakes, e.g.,Nikolaou et al. 2002) because of their high volatility;however, they are the most common groundwatercontaminants (Golfinopoulos et al. 2001). In the north-east of the landfill, surface water samples from ZC,located in a sewage ditch, yielded similar results overtwo sampling periods. MAHs accounted for 78 % ofthe total organic compounds, CAHs for 19 %, andPAHs for <2 %. In the pond surface water sampleZD, near the southeast of the landfill, fewer CAHswere detected (only tetrachloroethylene, 0.22 μg/L) inthe summer period, with no organic pollutantsdetected in the winter. In the southwest of the landfill,the leachate sample ZE showed CAHs to account for50 % of the total organic compounds, MAHs for 48 %,PAHs for <3 %, and OCPs for only 0.4 %. Figure 12shows that PAH concentrations of groundwater char-acterized by a declining trend with increasing distancefrom the landfill center. According to the exceedanceof 1 μg/L, the contaminated distance can be estimatedto be 1,200 m from this map. Pollutant concentrationsvary greatly in different directions (Fig. 13). Towardsthe east and south direction, the concentration decreasesvery significantly away from the landfill. By contrast,there is no obvious decreasing trend of the pollutantconcentrations in the north direction. Combined withthe variation of the different chemical composition andpollutants in shallow aquifers (Fig. 5a–f), it indicates


that leakages from the landfill into the aquifer havedeveloped in different directions and to different extents.Generally, as the groundwater table rises in the winterseason, pH lowers and the concentration of organiccompounds (especially PAHs) increases, compared tothose in the summer season. Large fluctuations in PAHsand BTEX concentrations in individual boreholes wereshown to be largely attributable to seasonal groundwaterflow variations, which can affect microbial and chemi-cal reaction that organic pollutants undergo and thetransport of gases and solutes through aquifer.

The sediments towards the north of the landfill,belonging to the southern margin of the YellowRiver paleo-alluvial fan, are composed of coarsegrains, including medium-fine grained sand and areknown to bear better quality water. This groundwaterhas been utilized for vegetable planting by local farm-ers, and as a result of groundwater exploitation, thenatural flow field has changed, causing wastewaterfrom the landfill to flow towards the north. The lowerchloride concentrations (Fig. 5a) of groundwater insouth portions of the aquifer relative to the leachateplume are consistent with mixing of leachate with

uncontaminated groundwater. The natural attenuationand/or mixing with fresh water can be shown from thechloride variation. In contrast, the sediments in thesouth of the landfill belong to the bank of HuaiheRiver and are characterized by fine grains, includingsilty clay and clay, which can prevent contaminationfrom spreading due to lower permeability. The farmersin the south of the landfill must plant wheat becausethere is less groundwater available for irrigation. As aresult, there is little groundwater abstraction. Localanthropogenic pollution input may have caused thehigh NO3

− and SO42− in east–south of the study area

(such as DW18). In the western part of the study area,groundwater discharges into the Jialu River. The dis-tribution characteristics of PAHs and BTEX are dif-ferent from other inorganic chemical composition. Thehighest concentration of PAHs is distributed some300 m SE of the landfill, while BTEX is highest atthe NW corner of the landfill.

Only CAHs were detected in the river water sam-ples (SUJL and SUY) in May 2009. This indicates thatriver water can become contaminated during the sum-mer period. CAHs, MAHs, and OCPs, with the excep-tion of PAHs, were detected in the sewage watersamples (SULG and SULD) to different extents duringthe multiple sampling campaigns. Lower detectionfrequencies of CAHs, MAHs, and OCPs and higherPAHs concentrations in groundwater than that in riverand sewage ditch suggest that there is no inflow ofpolluted groundwater into the stream, and the contam-inated river and waste water from sewage ditch aregenerally not recharging the groundwater body. Thatis, the interaction between groundwater and the sur-face water is likely weak. The municipal waste wateremission and the agricultural irrigation return flowinto the river and the sewage ditch could be sourcesof surface water pollution upstream; however, the

Fig. 13 Variation of different contaminants in different directions

Fig. 12 Variations of PAHs concentrations with distances fromthe landfill center


landfill body is the main pollution source for ground-water contamination in the study area.

Vertical distribution of organic pollution

In the contaminated area (dashed line, Fig. 1), thewaste dump is immersed in groundwater. A ground-water mound has formed at the landfill site and dif-fused into the surroundings. Local water tablegradients just below and around the landfill may differfrom the general gradients because the landfill mayhave a different hydrogeology than the surroundingstrata (Christensen 1992). The unlined Zhoukou land-fill without any top cover may result in a larger infil-tration than in the surrounding soil, and if leachate isnot removed quickly, this may result in a local watertable mound potentially affecting the local gradients.Local mounding effects are enhanced lateral spreadingof the leachate plume and to downward directed hy-draulic gradients in the groundwater zone beneath thelandfill. The enhanced lateral spreading of the plumemay increase the volume of contaminated groundwa-ter and its spatial extent, but provides increased dilu-tion of contaminants.

Controlled by aquifer heterogeneity, contaminatedgroundwater has reached the shallow aquifer to depthsof 13–25 m, which is composed of fine and middlesand with hydraulic conductivity of 11–12 m/day (Qu2010). The results show that, near the landfill, theshallow aquifer within 25 m depth has been contami-nated, but not the deeper aquifer at 50 m depth. The

deepest well, ZK1 at 300 m depth, was not contami-nated by organic pollutants, with the exception of0.44 μg/L toluene, which may have resulted fromregional groundwater flow. Additionally, water sam-ples from the monitoring wells (such as ZKW andZ16) near the gas station (Fig. 11b) were not detectedto have higher Phe and Nap, which are plentiful infresh fuels (Colombo et al. 2005a, b; Iturbe et al.2005). Therefore, their nonprevalence in groundwaterseems to show no recent leaks of petroleum productsfrom tanks and pipelines.

The main conclusion can be drawn that based onthe organic contaminants distribution of different wa-ter bodies, the interaction between groundwater andthe surface water (including water from the river,gully, and sewage ditch) is probably weak in this area.The landfill body is hence the main pollution sourcefor groundwater contamination. However, Municipalwaste water emissions and agricultural irrigation re-turn flow into the river and the sewage ditch could bethe source of surface water pollution upstream.

Quality assessment and conceptual model

River water and groundwater monitoring together withthe analysis of the organic pollutants demonstratedthat the landfill leachate is significantly impacting onthe surrounding aquatic environments. PAHs are themajor pollutants in groundwater surrounding theZhoukou landfill and can be used to evaluate ground-water quality. Compared with PAHs concentrations

Table 9 PAHs concentrations in surface water reported in the world

Year N Range(ng/L) Mean±SD(ng/L) References

River water:

Jialu and Ying River December 2009 6 nd–10 60±107.3 This study

Hai River, Tianjin, China 16 115±58.2 Shi et al. 2005

Tonghui River, Beijing, China 2002.Apr 16 193–2,651 762±777 Zhang et al. 2004

Middle and lower Yellow River, China 2004.Jun 15 179–369 248±78 Li et al. 2006

Xihe River, China 2006,Aug. 11 26–384 151±22 Song et al.2007

Gaoping River, Taiwan, China 1999/2000 16 10–9,400 430 Doong and Lin 2004

Lower Mississippi River, USA 1999 13 5.6–68.9 40.8±32.9 Mitra and Bianchi 2003

Elbe River, Hamburg, Germany 1992/1993 16 107–124 116±12 Götz et al.1998

Lower Seine River, France 1993.Oct 11 4–36 20±13 Fernandes et al. 1997

Lower Brisbane River, Australia 2001/2002 15 5–12 8.2±3.0 Shaw et al. 2004

Malaysian River, Malaysia 2009 3,925–5,126 4,682±238 Geik et al. 2009

N number of PAH compounds analysed in each study


reported for other contaminated rivers in China, USA,Germany, Australia, and other countries (Table 9),

Jialu and Ying river water in this study area are char-acterized by relatively lower PAHs concentrations

Fig. 14 Conceptual diagram of leachate migration in the Qua-ternary aquifers surrounding the Zhoukou landfill. The data ofgeological background is referenced from Qu (2010). Thegroundwater zone contaminated by leachate is determined by

the distribution of the total PAHs concentrations in groundwatersamples. The boundary line of the contaminated zone is thecontour of total PAHs concentration 0.1 μg/L, which is themaximum permissible value of the WHO standard


(<0.01 μg/L), indicating light organic contaminationin river water. The groundwater close to the river bedshould be diluted with lower PAHs concentrations ifthis was a major recharge source in the wet season.However, Z16 is featured by higher PAHs concentra-tion (1.07 μg/L) in May 2009 than that (0.33 μg/L) inDecember 2009, suggesting that the groundwater re-charge from river water is horizontally very limitedwithin short distances and that the dominant driver ofgroundwater flow and contaminant migration is thegroundwater mound surrounding the landfill.

Figure 14 shows the conceptual model of leachatemigration in the Quaternary aquifers surrounding theZhoukou landfill. There is a groundwater mound dif-fused into the surroundings at the landfill site. The levelsof PAHs were generally higher in the vicinity of thelandfill. Based on the distribution of PAHs concentra-tions in groundwater along the NS and WE hydrogeo-logical cross-sections (Fig. 14a-b), the groundwaterzone contaminated by leachate can be circled by thecontour of total PAHs concentration 0.1 μg/L, which isthe maximum permissible value of the WHO standard.The results suggest that groundwater beneath theZhoukou landfill and within 50 m depth is not suitableas a drinking water source, and pollution control shouldbe improved and enhanced in this area (for example,with the construction of artificial liner or providingimpermeable clay cover to reduce water infiltration intothe waste site). The groundwater contaminated zonevaries from May to December. From Fig. 14, it can beseen that the zone becomes smaller in theW–E directionbut has little change in the N–S direction between thesetimes, suggesting anisotropy in the local permeabilitydistribution.

Due to the complexity of leachate migrationthrough landfills, fundamental aspects of subsurfacecontaminant transport include the thickness of theunsaturated zone, the permeability and moisture con-tent of the earth materials within the unsaturated zone,and the hydraulic conductivity and local hydraulicgradient of geological units in the saturated zone(Taylor and Allen 2006). Poorly conductive units un-derlying the landfill, e.g., clay-rich material or thepresence of an installed artificial liner can reduceleachate migration. On the other hand, discontinuitiesof the landfill bottom such as fissures and joints in thesubsurface or faults or holes in a liner, dramaticallyincrease leachate flow. Access to hydrogeological in-formation is thus vital for situation assessments and

designing lining systems both beneath and down-stream of landfills. With the gradual expansion of theZhoukou city area, the landfill has been surrounded byurban planning area. If the local government does nottake preventive measures, the existing waste will con-tinue long-term groundwater contamination. Long-term detailed monitoring programs are essential todevelop conceptual models of natural attenuation,and studies need to allow the recognition that ourunderstanding of microbial transformation pathwaysis constantly changing.

Conclusions and environmental implications

The investigation of organic contamination aroundZhoukou landfill shows the present status of pollutionin sediments and surface and groundwater. This paperhas provided important data on parent PAH levels andother organic contaminants in the water and sedimentsof the Zhoukou landfill in Henan Province, China.Some conclusions can be drawn as follows:

1. The main source and pathway for organic contam-ination is infiltration of rainfall in the vicinity ofthe landfill, which has created a local groundwatermound.

2. Detected organic contaminants include MAHs,CAHs, OCPs, and PAHs. The concentrations ofthese compounds are affected by seasonal ground-water table fluctuations. PAHs are the main organ-ic contaminant in this study area. Among thedetected eleven PAHs, Phe, Flu, and FLT identi-fied by PCA are the three dominant in most of thegroundwater samples. PAH diagonostic ratios, in-cluding FLT/Pyr, FLT/(FLT + Pyr), Phe/Ant, andCHR/BaA, suggest the mixture of petrogenic andpyrolytic contaminations of groundwater near theZhoukou landfill.

3. Higher NO3− concentrations and lower pH values

under aerobic conditions enhance degradation ofPAHs in December, resulting in lower PAHs andHCO3

- concentrations.4. The organic contaminants detected from different

water bodies show that the interaction betweengroundwater and the surface water (including wa-ter from the river, gully, and sewage ditch) is weakin this area. The landfill body is the main pollutionsource for groundwater contamination in this


study area. The municipal waste water emissionand the agricultural irrigation return flow into theriver and the sewage ditch could be sources ofsurface water pollution upstream.

5. Based on the PAHs concentrations distribution, aconceptual model of leachate migration in theQuaternary aquifers surrounding the Zhoukoulandfill has been developed to describe the con-tamination processes. The groundwater zone con-taminated by leachate has been identifiedsurrounding the landfill. An oval-shaped pollutionhalo has formed, spanning 3 km from west to eastand 2 km from south to north, and mainly occursin the area with groundwater level depths of 2–4 m. High detection rates of contaminants (espe-cially PAHs) in groundwater from the shallowaquifer at 18–30 m indicate that it has been heavi-ly contaminated. The deeper aquifer at depths>50 m has not yet been contaminated by organicpollutants. The existence of clay and silty claylayers with stable thicknesses at about 20 m (30–50 m depth) may contribute to these results.

In order to comprehensively evaluate groundwaterquality and protect drinking water, it is important toinvestigate spatial and temporal distributions of organ-ic pollution in the subsurface environment. The find-ings point to an urgent need to establish a robustmonitoring procedure for persistent organic pollutantssuch as PAHs, not only in water bodies and sedimentsbut also in the relationships between microorganisms,and in regards to aquifer physico-chemical parameters.Any exceedance in organic pollutant concentrationsover the environmental quality standards should berapidly reported and the necessary actions taken tomitigate the effects. Additionally, it is necessary toimprove urban sewage facilities, such as establishingunderground sewage pipes and implementing antisee-page treatments along the sewage ditches, in order toreduce leakage of landfill leachate.

Acknowledgments This research was financially supported bythe Exploratory Forefront Project (no. 2012QY007) for the StrategicScience Plan in the Institute of Geographic Sciences and NaturalResources Research, Chinese Academy of Sciences, and was un-dertaken as part of a groundwater survey project titled “Investiga-tion and evaluation of typical contaminated sites in Zhoukou regionof Huaihe River Plain” (no. 1212010634505). The authors aregrateful to Mr. Xie Shiyong and Qu Zewei from School of Envi-ronmental Studies, China University of Geosciences, for their helpand support during water sampling in the field and data collection.

References

Akyüz, M., & Çabuk, H. (2010). Gaseparticle partitioning andseasonal variation of polycyclic aromatic hydrocarbons inthe atmosphere of Zonguldak, Turkey. Science of the TotalEnvironment, 408, 5550–5558.

APHA (1992). Standard methods for the examination of waterand wastewater, 18th edn. Washington: American PublicHealth Association.

Basberg, L., Banks, D., & Saether, O. M. (1998). Redox pro-cesses in groundwater impacted by landfill leachate.Aquatic Geochemistry, 4, 253–272.

Baudoin, C., Charveron, M., Tarroux, R., & Gall, Y. (2002).Environmental pollutants and skin cancer. Cell Biology andToxicology, 18, 341–348.

Baumard, P., Budzinski, H., & Garrigues, P. (1998). Polycyclicaromatic hydrocarbons in sediments and mussels of thewestern Mediterranean Sea. Environmental Toxicologyand Chemistry, 17, 765–776.

Baumard, P., Budzinski, H., Michon, Q., Garrigues, P., Burgeot,T., & Bellocq, J. (1998). Origin and bioavailability ofPAHs in the Mediterranean Sea from mussel and sedimentrecords. Estuarine Coastal and Shelf Science, 47, 77–90.

Borden, R. C., Gomez, C. A., & Becker, M. T. (1995).Geochemical indicators of intrinsic bioremediation.Ground Water, 33, 180–189.

Bosma, T. N. P., Harms, H., & Zehnder, J. B. (2001)Biodegradation of xenobiotics in the environment andtechnosphere. In B. Beek (Ed.), The handbook of environ-mental chemistry, Park K: biodegradation and persistence,vol. 63. Berlin: Springer.

Brigmon, R. L., Camper, D., & Stutzenberger, F. (2002).Bioremediation of compounds hazardous to health andthe environment—an overview. In V. P. Singh & R. D.Stapleton (Eds.), Biotransformations: bioremediation tech-nology for health and environmental protection (pp. 1–28).Amsterdam: Elsevier Science.

Budziński, H., Jones, I., Bellocq, J., Pierrad, C., & Garrigues, P.(1997). Evaluation of sediment contamination by polycy-clic aromatic hydrocarbons in the Gironde estuary. MarineChemistry, 58, 85–97.

Cao, Z. H., Wang, Y. Q., Ma, Y. M., Xu, Z., Shi, G. L., Zhuang,Y. Y., et al. (2005). Occurrence and distribution of polycy-clic aromatic hydrocarbons in reclaimed water and surfacewater of Tianjin, China. Journal of Hazardous Materials,A122, 51–59.

Chapman, D. (1992). Water quality assessments. A guide to theuse of biota, sediments and water in environmental moni-toring (pp. 371–460). London, UNESCO/WHO/UNEP,Chapman & Hall.

Christensen, T. H. (1992). Attenuation of leachate pollutants ingroundwater, landfilling of waste. In T. H. Christensen, R.Cossu, & R. Stegmann (Eds.), Leachate (pp. 441–469).London: Elsevier Applied Science.

Christensen, T. H., Kjeldsen, P., Bjerg, P. L., Jensen, D. L.,Ch r i s t en sen , J . B . , Baun , A . , e t a l . ( 2001 ) .Biogeochemistry of landfill leachate plumes. AppliedGeochemistry, 16, 659–718.

Colombo, J. C., Barreda, A., Bilos, C., Cappelletti, N.,Demichelis, S., & Lombardi, P. (2005a). Oil spill in the


Rio de La Plata estuary, Argentina: 1. Biogeochemicalassessment of waters, sediments, soils and biota.Environmental Pollution, 134, 277–289.

Colombo, J. C., Barreda, A., Bilos, C., Cappelletti, N., Migoya,M. C., & Skorupka, C. (2005b). Oil spill in the Rio de LaPlata estuary, Argentina: 2-hydrocarbon disappearancerates in sediments and soils. Environmental Pollution,134, 267–276.

Cui, Y. C., & Fu, T. (1998). Harm of water pollution and organicpollutants in drinking water resources in China. UrbanEnvironment &Urban Ecology (in Chinese), 11(3), 23–25.

Daniel, R.A., & Borden, R.C. (1997). Spatial variability inintrinsic bioremediation rates: effects on contaminant trans-port. In: B.C. Alleman, & A. Leeson (Eds.), In situ and on-site bioremediation, (Vol. 1, pp. 29–34). Columbus:Battelle.

Davis, G. B., Barber, C., Power, T. R., Thierrin, J., Patterson, B.M., Rayner, J. L., et al. (1999). The variability and intrinsicremediation of a BTEX plume in anaerobic sulphate-richgroundwater. Journal of Contaminant Hydrology, 36, 265–290.

De La Torre-Roche, R. J., Lee, W. Y., & Campos-Díaz, S. I.(2009). Soil-borne polycyclic aromatic hydrocarbons in ElPaso, Texas: analysis of a potential problem in the UnitedStates/Mexico border region. Journal of HazardousMaterials, 163, 946–958.

De Luca, G., Furesi, A., Leardi, R., Giovanni, M. G., Panzanelli,A., Costantina, P. P., et al. (2004). Polycyclic aromatichydrocarbons assessment in the sediments of the PortoTorres Harbor (Northern Sardinia, Italy). MarineChemistry, 86, 15–32.

Doong, R. A., & Lin, Y. T. (2004). Characterization and distri-bution of polycyclic aromatic hydrocarbon contaminationsin surface sediment and water from Gao-ping River,Taiwan. Water Research, 38, 1733–1744.

Dyreborg, S., & Arvin, E. (1995). Inhibition of nitrification bycreosote contaminated water. Water Research, 29, 1603–1606.

Eggen, T., Moeder, M., & Arukwe, A. (2010). Municipal land-fill leachates: a significant source for new and emergingpollutants. Science of the Total Environment, 408, 5147–5157.

Eichelberg, J.W., & Bundle, W.L. (1989). US measurement ofpurgeable organic compounds in water by capillary columngas chromatography/mass spectrometry. EPA method524.2 revision 3.0. Columbus: EPA.

Essaid, H. I., Bekins, B. A., Godsy, E. M., Waren, E.,Baedecker, M. J., & Cozzarelli, I. M. (1995). Simulationof aerobic and anaerobic biodegradation processes at acrude oil spill site. Water Resources Research, 31(12),3309–3327.

Fernandes, M. B., Sicre, M. A., Boireau, A., & Tronczynski, J.(1997). Polyaromatic hydrocarbon (PAH) distributions inthe Seine River and its Estuary. Marine Pollution Bulletin,34, 857–867.

Geik, K.H., Zakaria, M.P., Lee, W.Y., & Haye, R., (2009).Landfill leachate as a source of polycyclic aromatic hydro-carbons PAHs to Malaysian waters. PPT-presentationonline. http://landbase.hq.unu.edu

Gogou, A., Apostolaki, M., & Stephanou, E. G. (1998).Determination of organic molecular markers in marine

aerosols and sediments: one-step flash chromatographycompound class fractionation and capillary gas chromato-graphic analysis. Journal of Chromatography A, 799, 215–231.

Golfinopoulos, S. K., Lekkas, T. D., & Nikolau, A. D. (2001).Comparison of methods for determination of volatile or-ganic compounds in drinking water. Chemosphere, 45,275–284.

Götz, R., Bauer, O. H., Friesel, P., & Roch, K. (1998). Organictrace compounds in the water of the River Elbe nearHamburg. Chemosphere, 36, 2103–2118.

Grbic-Galic, D., & Vogel, T. M. (1987). Transformation oftoluene and benzene by mixed methanogenic cultures.Applied and Environmental Microbiology, 53, 254–260.

Gschwend, P. M., & Hites, R. A. (1981). Fluxes of polycyclicaromatic hydrocarbons to marine and lacustrine sedimentsin the Northeastern US. Geochimica et CosmochimicaActa, 45, 2359–2367.

Haritash, A. K., & Kaushik, C. P. (2009). Biodegradationaspects of polycyclic aromatic hydrocarbons (PAHs): areview. Journal of Hazardous Materials, 169, 1–15.

Hinga, K. R. (2003). Degradation rates of low molecular weightPAH correlate with sediment TOC in marine subtidal sedi-ments. Marine Pollution Bulletin, 46, 466–474.

Iturbe, R., Flores, C., Flores, R. M., & Torres, L. G. (2005).Subsoil TPH and other petroleum fractions-contaminationlevels in an oil storage and distribution station in north-central Mexico. Chemosphere, 61, 1618–1631.

Keener, W. K., & Arp, D. J. (1994). Transformation of aromaticcompounds by Nitrosomonas europaea. Applied andEnvironmental Microbiology, 4, 1914–1920.

Lee, B. K., & Lee, C. B. (2004). Development of an improveddry and wet deposition collector and the atmospheric de-position of PAHs onto Ulsan Bay Korea. AtmosphericEnvironment, 38, 863–871.

Lee, J. Y., Cheon, J. Y., Lee, K. K., Lee, S. Y., & Lee, M. H.(2001). Statistical evaluation of geochemical parameterdistribution in a ground water system contaminated withpetroleum hydrocarbons. Journal of EnvironmentalQuality, 35, 1548–1563.

Li, G. C., Xia, X. H., Yang, Z. F., Wang, R., & Voulvoulis, N.(2006). Distribution and sources of polycyclic aromatichydrocarbons in the middle and lower reaches of theYellow River, China. Environmental Pollution, 144, 985–993.

Lu, G. P., Clement, T. P., Zheng, C. M., & Wiedemeier, T. H.(1999). Natural attenuation of BTEX compounds: modeldevelopment and field-scale application. Ground Water, 37(5), 707–717.

Lyngkilde, J., & Christensen, T. H. (1992). Redox zones of alandfill leachate pollution plume (Vejen, Denmark).Journal of Contaminant Hydrology, 10, 273–289.

Magi, E., Bianco, R., Ianni, C., & Di Carro, M. (2002).Distribution of polycyclic aromatic hydrocarbons in thesediments of the Adriatic Sea. Environmental Pollution,119, 91–98.

Maliszewska-Kordybach, B. (1996). Polycyclic aromatichydrocarbons in agricultural soils in Poland: prelimi-nary proposals for criteria to evaluate the level of soilcontamination. Applied Geochemistry, 11, 121–127.


McNally, D. L., Mihelcic, J. R., & Lueking, D. R. (1999).Biodegradation of mixtures of polycyclic aromatic hydro-carbons under aerobic and nitrate-reducing conditions.Chemosphere, 38(6), 1313–1321.

Meuller, J. G., Chapman, P. J., & Pritchard, P. H. (1989). Actionof fluoranthene-utilizing community on polycyclic aromat-ic hydrocarbon components of creosote. Applied andEnvironmental Microbiology, 55, 3085–3090.

Mitra, S., & Bianchi, T. S. (2003). A preliminary assessment ofpolycyclic aromatic hydrocarbon distributions in the lowerMississippi River and Gulf of Mexico. Marine Chemistry,82, 273–288.

Neilson, A. H., & Allard, A. S. (1998). Microbial metabolism ofPAHs and heteroarenes. In A. H. Neilson (Ed.), Thehandbook of environmental chemistry (pp. 2–80). Berlin:Springer. vol. 3, part I.

Nielsen, Per H., Bjarnadóttir, H., Winter, Pia L., & Christensen,T.H. (1995). In situ and laboratory studies on the fate ofspecific organic compounds in an anaerobic landfill leach-ate plume. 2. Fate of aromatic and chlorinated aliphaticcompounds. Journal of Contaminant Hydrology, 20, 51–66.

Nikolaou, A. D., Golfinopoulos, S. K., Kostopoulou, M. N.,Kolokytas, G. A., & Lekkas, T. D. (2002). Determinationof volatile organic compounds in surface waters and treatedwastewater in Greece. Water Research, 36, 2883–2890.

Persson, L., Alsberg, T., Ledin, A., & Odham, G. (2006).Transformations of dissolved organic matter in a landfillleachate—a size exclusion chromatography/mass spectro-metric approach. Chemosphere, 64, 1093–1099.

Pies, C., Hoffmann, B., Petrowsky, J., Yang, Y., Ternes, T. A., &Hofmann, T. (2008). Characterization and source identifi-cation of polycyclic aromatic hydrocarbons (PAHs) in riverbank soils. Chemosphere, 72, 1594–1601.

Postma, D., Boesen, C., Kristiansen, H., & Larsen, F. (1991).Nitrate reduction in an unconfined sandy aquifer—waterchemistry, reduction processes, and geochemical modeling.Water Resources Research, 27(8), 2027–2045.

Qu, Z.W. (2010). Hydrogeological conditions and evaluation ofgroundwater resources in Yingbei region of Zhoukou city(pp. 12–13). Master degree thesis in China University ofGeosciences (Wuhan), (in Chinese with English abstract).

Rainwater, K., Mayfield, M. P., Heintz, C., & Claborn, B. J.(1993). Enhanced in situ biodegradation of diesel fuel bycyclic vertical water table movement: preliminary studies.Water Environment Research, 65, 717–725.

Ravindra, K., Sokhi, R., & van Grieken, R. (2008a).Atmospheric polycyclic aromatic hydrocarbons: source at-tribution, emission factors and regulation. AtmosphericEnvironment, 42, 2895–2921.

Ravindra, K., Wauters, E., & van Grieken, R. (2008b). Variationin particulate PAHs levels and their relation with the trans-boundary movement of the air masses. Science of the TotalEnvironment, 396, 100–110.

Rothermich, M. M., Hayes, L. A., & Lovley, D. R. (2002).Anaerobic, sulfate-dependent degradation of polycyclic ar-omatic hydrocarbons in petroleum-contaminated harborsediment. Environmental Science & Technology, 36,4811–4817.

Rügge, K., Bjerg, P. L., & Christensen, T. H. (1995).Distribution of organic compounds from municipal solid

waste in the groundwater downgradient of a landfill(Grindsted, Denmark). Environmental Science &Technology, 29, 1395–1400.

Sahu, S. K., Pandit, G. G., & Sadasivan, S. (2004). Precipitationscavenging of polycyclic aromatic hydrocarbons inMumbai, India. Science of the Total Environment, 318,245–249.

Shaw, M., Tibbetts, I. R., & Jochen, M. F. (2004). MonitoringPAHs in the Brisbane River and Moreton Bay, Australia,using semipermeable membrane devices and EROD activ-ity in yellowfin bream, Acanthopagrus australis.Chemosphere, 56, 237–246.

Shi, Z., Tao, S., Pan, B., Fan, W., He, X. C., Zuo, Q., et al.(2005). Contamination of rivers in Tianjin, China by poly-cyclic aromatic hydrocarbons. Environmental Pollution,134, 97–111.

Sicre, M. A., Marty, J. C., Saliot, A., Aparicio, X., Grimalt, J., &Albaiges, J. (1987). Aliphatic and aromatic hydrocarbonsin different sized aerosols over the Mediterranean Sea:occurrence and origin. Atmospheric Environment, 21,2247–2259.

Sinke, A. J. C., Dury, O., & Zobrist, J. (1998). Effects of afluctuating water table: column study on redox dynamicsand fate of some organic pollutants. Journal ofContaminant Hydrology, 33, 231–246.

Soclo, H. H., Garrigues, P., & Ewald, M. (2000). Origin ofpolycyclic aromatic hydrocarbons (PAHs) in coastal ma-rine sediments: case studies in Cotonou (Benin) andAquitaine (France) areas. Marine Pollution Bulletin, 40,387–396.

Song, X. Y., Sun, L. N., Wang, X., Li, X. X., & Sun, T. Y.(2007). Organic contamination status of Xihe River surfacewater and its riverbank underground water. ChineseJournal of Ecology (in Chinese with English abstract), 26(12), 2057–2061.

Stafford, D. A. (1974). The effect of phenols and heterocyclicbases on nitrification in activated sludges. Journal ofApplied Bacteriology, 37, 75–82.

Taylor, R., & Allen, A. (2006). Waste disposal and landfill:Information needs. In: O. Schmoll, G. Howard, J.Chilton, & I. Chorus (Eds.). Protecting groundwater forhealth: managing the quality of drinking-water sources(Chapter 12, pp. 1–12). London: IWA.

Tobiszewski, M., & Namieśnik, J. (2012). PAH diagnostic ratiosfor the identification of pollution emission sources.Environmental Pollution, 162, 110–119.

Tong C.S. (2012). The characterization of groundwaterpollution and prevention measures around an old land-fill: heavy metals and polycyclic aromatic hydrocar-bons (pp. 22–31). Ph.D. degree thesis in ChinaUniversity of Geosciences (Wuhan) (in Chinese withEnglish abstract).

van Breukelen, B.M. (2003). Natural attenuation of landfillleachate: a combined biogeochemical process analysisand microbial ecology approach (pp. 105–112). Doctoralthesis Vrije Universiteit Amsterdam.

van Brummelen, T. C., van Hattum, B., Crommentuijn, T.& Kalf, D. E. (1998). Bioavailability and ecotoxicol-ogy of PAHs. In A. H. Neilson (Ed.), PAHs andrelated compounds chemistry (pp. 175–215). NewYork: Springer.


Wan, X. L., Chen, J. W., Tian, F. L., Sun, W. J., Yang, F. L., &Saiki, K. (2006). Source apportionment of PAHs in atmo-spheric particulates of Dalian: factor analysis with nonneg-ative constraints and emission inventory analysis.Atmospheric Environment, 40, 6666–6675.

Wang, Y., Li, P. H., Li, H. L., Liu, X. H., & Wang, W. X. (2010).PAHs distribution in precipitation at Mount Taishan,China: identification of sources and meteorological influ-ences. Atmospheric Research, 95, 1–7.

WHO (2006). Guidelines for drinking-water quality, 3rd edn, vol.1. Recommendations. Geneva: World Health Organization.

Wiedemeier, T., Wilson, J.T., Kampbell, D.H., Miller, R.N., &Hansen, J.E. (1995). Technical protocol for implementingintrinsic remediation with long-term monitoring for naturalattenuation of fuel contamination dissolved in groundwa-ter, Vols. I and II. Air Force Centre for EnvironmentalExcellence AFCEE, Technology Transfer Division,Brooks AFB, San Antonio, TX.

Wilson, R. D., Thornton, S. F., & Mackay, D. M. (2004).Challenges in monitoring the natural attenuation of spatial-ly variable plumes. Biodegradation, 15, 359–369.

Yunker, M. B., Macdonald, R. W., Vingarzan, R., Mitchell, R.H., Goyette, D., & Sylvestre, S. (2002). PAHs in the Fraser

River basin: a critical appraisal of PAH ratios as indicatorsof PAH source and composition. Organic Geochemistry,33, 489–515.

Zaidi, B. R., & Imam, S. H. (1999). Factors affecting microbialdegradation of polycyclic aromatic hydrocarbon phenan-threne in the Caribbean coastal water. Marine PollutionBulletin, 38(8), 737–742.

Zakaria, M. P., Takada, H., Tsutsumi, S., Ohno, K., Yamada, J.,Kouno, E., et al. (2002). Distribution of polycyclic aromatichydrocarbons (PAHs) in rivers and estuargenic PAHs.Environmental Science & Technology, 36, 1907–1918.

Zhang, Z. L., Huang, J., Yu, G., & Hong, H. S. (2004).Occurrence of PAHs, PCBs and organochlorine pesticidesin the Tonghui River of Beijing, China. EnvironmentalPollution, 130, 249–261.

Zhang, W., Zhang, S., Wan, C., Yue, D., Ye, Y., & Wang, X.(2008). Source diagnostics of polycyclic aromatic hydro-carbons in urban road runoff, dust, rain and canopythroughfall. Environmental Pollution, 153, 594–601.

Zhou, J. L., & Maskaoui, K. (2003). Distribution of polycyclicaromatic hydrocarbons in water and surface sedimentsfrom Daya Bay, China. Environmental Pollution, 121,269–281.


evaluation of organic contamination in urban groundwater surrounding...

Documents