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Evaluation of the impact of an uncontrolled land ll on surroundinggroundwater quality, Zhoukou, China

Dongmei Han a , , Xiaoxia Tong b , Matthew J. Currell c, Guoliang Cao d , Menggui Jin e , Changshui Tong e

a Key Laboratory of Water Cycle & Related Land Surface Processes, Institute of Geographic Sciences and Natural Resources Research, Chinese Academy of Sciences, Beijing 100101, Chinab School of Water Resources and Environment, China University of Geosciences, Beijing 100083, Chinac School of Civil, Environmental and Chemical Engineering, RMIT University, Melbourne, Australiad Center for Water Research, College of Engineering, Peking University, Beijing 100871, Chinae School of Environmental Studies, China University of Geosciences, Wuhan 430074, China

a b s t r a c ta r t i c l e i n f o

Article history:Received 1 April 2012Accepted 30 September 2013Available online 9 October 2013

Keywords:Uncontrolled land llInorganic contaminationGroundwater hydrochemistryGroundwater qualityContaminant transport model

Our groundwater pollution investigation of an uncontrolled municipal land ll aims to better understand the possi-ble impact of leachate percolation on groundwater quality.The study examinedgroundwater inorganic contamina-tion arounda municipalland ll site in Zhoukou city, Henan province, China.Stable isotopic compositions ( δ18 O andδ2H) and concentrations of various physico-chemical parameters were determined in surface and groundwatersamples collected from the study area. δ18 O and δ2H compositions reveal that elevation effects and/or evaporativeprocesses occur to various extents prior to water reaching the saturated zone. There is no serious heavy metalcontamination in this area. Principal component analysis was applied, and identi ed TDS, Cl− , NH4

+ , Fe and Mnconcentrations as the mainindicators of groundwater pollution caused by land ll leachatepercolation.The ground-water quality of shallow aquifer is likely dominated by irrigation return ow in the summer. A two-dimensionaladvective – dispersive transport model was established using MODFLOW and MT3DMS to explore the controls oncontaminant transport. Chloride transport simulations with steady state and transient ow models show that thecontamination plume is still constrained in the upper aquifer after 13yearsof operation of the land ll. The contrastbetween the simulated (relatively short) migration distance of the chloride plume, and the widely observed highchloride concentrations in the upperaquifer indicate other possible local contributions to the salinityof the aquifer.Thetransient model indicatesthat seasonaldynamics, such as temporalwater level variations,may have a great in-

uence on the migration of the plume beneath the land ll by controlling verticalhydraulic gradients and uxes. Ingeneral, theshallowgroundwater(within 30m depth) aroundthe Zhoukou land ll is notsuitablefor drinking, andpollution control should be improved and enhanced in this area.

© 2013 Elsevier B.V. All rights reserved.

1. Introduction

The common land ll receives a mixture of municipal, commer-cial and mixed industrial waste, typically producing a wide rangeof pollutant compounds with resulting environmental, wildlife andhuman health impacts. A large number of land ll-caused ground-water pollution cases have been reported around the world ( Abu-Rukah and Al-Kofahi, 2001; Arneth et al., 1989; Christensen et al.,1998; Flyhammar, 1995; Looser et al., 1999; Rapti-Caputo andVaccaro, 2006; Saarela, 2003 ), and substantial resources spent ongroundwater remediation in these areas suggest that land ll

leachates, generated by excess rainfall percolating through wastedump, are signi cant sources of groundwater pollution and pose majorrisks to groundwater quality ( Christensen et al., 2001; Niininen et al.,1994 ). To date, there are few studies investigating theextent and mecha-nisms of groundwater contaminationin China.This is despitethefact thatit is known that contamination is almost ubiquitous in most urban areas,and that the nation faces a potential publichealth crisis dueto groundwa-ter pollution ( Li, 2013).

Many industries dispose off waste chemicals by placing them in un-lined soil pits and lagoons, where they can easily migrate to surface andgroundwater. Physical, chemical, and microbial processes within thewaste can be combined to transfer pollutants from the waste materialinto percolating rainwater. Furthermore, it has been shown that heavymetals (e.g. Cd, Cu, As, Pb, and Cr6+ ), ammonium (NH 4

+ ), and nitrate(NO3

− ) from uncontrolled land lls pose a major threat to human healththrough ingestion of contaminated groundwater( Lee et al., 2006 ). Mul-tiple indicators, including major hydrochemistry, and stable isotopes of

Journal of Geochemical Exploration 136 (2014) 24 – 39

Corresponding author.E-mail address: dmeihan@gmail.com (D. Han).

0375-6742/$ – see front matter © 2013 Elsevier B.V. All rights reserved.

http://dx.doi.org/10.1016/j.gexplo.2013.09.008

Contents lists available at ScienceDirect

Journal of Geochemical Exploration

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water ( δ18 O,δ2H),have been usedto providean improved understandingof the complex hydraulic and hydrological conditions present in urbanareas ( Barrett et al., 1999; Foster, 1998; Morris et al., 2006; Osenbrücket al., 2007 ).

Of the seven large rivers in China, the Huaihe River has the highestpopulation density and the most serious water contamination issues(Cui and Fu, 1998 ). With rapid economic development in many citiesof the Huaihe River Basin, increasing demand for water is con icting

withwater resource shortages, severely restrictingthe sustainable utiliza-tion of the region's water and the development of many local economies.

Zhoukou city is located in the southeast valley of the Huaihe River Basin,in the con uence of the Ying and the Jialu Rivers, which belong to tribu-taries of the Huaihe River (Fig. 1). Groundwater is the main source of water supply for industrial, agricultural and domestic use in Zhoukoucity (Zhao et al., 2010 ) and both urban industrial and domestic wasteand sewage discharge to surface water and groundwater. Groundwaterquality around Zhoukou city has been deteriorating rapidly over the lastfew decades. Drinking ungraded water and contaminated groundwater

long term can result in endemic diseases. Identi cation of the sources of contamination and understanding the pathways of contaminants to and

Fig. 1. Study area and location of sampling sites around the Zhoukou land ll. LEGEND: 1—residential area; 2 —land ll; 3—waters; 4 —groundwater monitoring wells (sampling wells withlabels in Table 1); 5 —Surface water sampling sites; 6 —shallow groundwater table contours (m a.s.l) from the monitoring data in December 2009; 7 —major groundwater- ow direction;

8— ow direction of surface water.

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Table 1Physico-chemical parameters of water samples around the land ll.

ID Sampling time Welldepth

(m)

Wellscreen(m)

Utilization T(°C)

pH EC(μ S/cm)

Eh(mV)

DO Ca2+ Mg2+ Na+ K+ Al Fe Mn

ZA Dec.2008 30 13.0– 25.2 Agricultural irrigation 16.6 7.71 964 256 6.38 122 45.2 30.2 6.5 0.07 0.2 0.33ZB Dec.2008 50 12.6– 25.1 Domestic well 14.6 7.7 1086 195 5.46 30.0 57.6 68.8 0.7 0.02 0.05 0.07ZC Dec.2008 – Sewage water 19.6 7.7 1602 − 208 4.03 69.1 35.9 179 16.9 0.3 0.4 0.11ZD Dec.2008 – Water from pond 13.9 8.14 638 138 3.06 44 27.4 62 6.9 0.3 0.5 0.13ZE Dec.2008 – Was te l eachate 11.1 7.93 3255 − 184 5.76 178 95.4 206 105 0.1 0.3 0.32ZF Dec.2008 18 13.2– 24.5 Agricultural irrigation 16.8 7.27 3365 328 1.54 259 68.7 125 3 1.1 1.1 0.17ZG Dec.2008 20 12.2 – 22.0 Domestic well 17.3 7.3 2235 283 4.97 218 68.8 114 3.6 0.9 0.8 0.05Z6 Dec.2008 20 13.0 – 24.0 Domestic well 17 7.3 1770 160 8.67 159 63.1 81.4 25.8 0.1 0.7 0.53Z6A Dec.2008 16 12.8 – 23.6 D omestic well 17 7.09 1734 149 5.59 1.55 58.7 103.6 91.1 0.01 0.01 0.31SW29 Dec.2008 9 5.2 – 9.0 Domestic well 16.3 7.49 1255 60 6.10 80.8 39.4 111.2 0.4 0.01 0.06 0.09SW49 Dec.2008 14 7.5 – 13.8 Domestic well 18.3 6.96 1231 137 0.48 275.6 83.5 106.8 27.6 0.07 0.3 0.01SW59 Dec.2008 6 4.2 – 6.0 Domestic well 17.5 7 .5 1325 − 5 0.54 124.2 49.0 93.0 0.1 0.2 0.2 0.45ZK1 Dec.2008 300 Urban water supply 22 8.47 893 72 2.41 28.5 18.2 148.5 1.2 0.01 0.01 0.04ZKE May 2009 27 11.7– 22.8 Observation well 51.7 65.1 225.0 3.1 0.02 0.05 1.04ZKW May 2009 27 12.6 – 25.3 O bservation well 18.3 8 .94 396 98 1.07 33.2 17.5 93.4 11.9 0 .32 0.19 0 .02ZKS May 2009 27 11.2– 25.1 Observation well 18.4 7.48 961 3 1.98 32.6 35.9 220 11.2 0.14 0.08 0.24ZKN May 2009 27 13.7– 24.1 O bservation well 19.3 6 .89 3644 115 1.92 128 210 369 55 0.07 0.05 0.67ZF May 2009 18 13.2 – 24.5 Agricultural irrigation 17.6 7.63 1316 151 2.55 76.8 81.6 196 6.1 0.01 0.05 0.16DW09 May 2009 16 9.2 – 15.9 Domestic well 18.9 7.39 1369 153 1.85 90.7 110 169 41.6 0.01 0.05 0.33DW18 May 2009 30 11.2 – 28.8 Domestic well 21.5 7.97 2930 106 2.05 16.6 46.7 255 4.1 0.01 0.05 0.1DW23 May 2009 18 7.6 – 17.8 Domestic well 18.4 7.52 1671 198 2.06 74.7 102 184 0.5 0.01 0.05 0.09

DW25 May 2009 30 8.2 – 25.7 Domestic well 20.5 7.5 1332 171 2.13 59.5 85.5 198 0.6 0.01 0.05 0.01ZG10 May 2009 9 4.5 – 8.8 Observation well 18.9 7.46 1076 73 2.29 58.1 37.8 234 0.3 0.05 0.05 2 .63ZG11 May 2009 9 4.8 – 9.0 Observation well 18.7 6.83 25 2.29 67.8 45.5 872 12.8 0.01 0.05 0.58ZG12 May 2009 9 5.2 – 8.9 Observation well 64.6 13.7 580 2 0.01 0.05 0.66Z16 May.2009 30 14.5 – 25.1 Agricultural irrigation 65.3 71.3 158 1.4 0.01 0.05 0.02Z34 May 2009 28 12.8 – 22.5 Agricultural irrigation 17.6 7.62 1809 194 2.35 123 96.3 202 0.6 0.01 0.05 0.32Z40 May 2009 30 13.2 – 26.3 Agricultural irrigation 17.2 7.61 1162 189 1.98 69 73.5 185 9.3 0.03 0.05 0.18ZC May 2009 – Sewage water 69.1 32.9 202 8.6 0.02 0.08 0.16SUJL May 2009 – River water 72.6 32.2 174 10.6 0.05 0.05 0.01SULG May.2009 – Sewage water 39.8 30.2 179 9.8 0.01 0.05 0.01SUY May 2009 – River water 74.9 37.6 182 10.7 0.24 0.15 0.01SULD May 2009 – Agricultural irrigation 81.3 53.1 175 6.6 0.15 0.23 0.29ZKE Dec. 2009 27 11.7– 22.8 Observation well 17.8 7.29 1676 − 91.3 161.0 55.3 161.0 1.1 0.6 9.3 3.92ZKW Dec. 2009 27 12.6– 25.3 Observation well 16.6 7.21 1652 193.6 75.6 61.0 8.4 0.42 1.1 0.73ZKS Dec. 2009 27 11.2– 25.1 Observation well 17.5 7.43 1353 − 87.2 92.0 37.5 150.0 3.1 0.82 11.1 2.98ZKN Dec. 2009 27 13.7– 24.1 Observation well 18.4 6.61 3999 397.6 212.6 346.0 61.6 0.05 1.8 1.46ZB Dec. 2009 50 12.6– 25.1 Domestic well 10.1 7.63 1164 50.7 104.0 61.1 72.8 0.9 0.02 0.1 0.14ZG9 Dec. 2009 9 4.3– 9.0 Observation well 18.5 7.39 1219 73.0 30.3 144.5 0.5 1.1 8.6 1.7Z06 Dec. 2009 15 7.5 – 14.6 Agricultural irrigation 16.8 7.25 1378 172.9 47.1 38.8 3.6 0.02 0.2 0.6Z40 Dec. 2009 30 13.2 – 26.3 Agricultural irrigation 16.5 7.33 1494 158.7 57.7 76.5 1.9 0.1 0.4 0.15Z16 Dec. 2009 30 14.5 – 25.1 Agricultural irrigation 17.4 7.73 787 102.4 27.9 34.3 8.6 0.8 4.8 1.35ZF Dec. 2009 18 13.2– 24.5 Agricultural irrigation 16.1 7.42 1502 134.9 54.0 119.5 9.7 0.14 0.8 0.56DW09 Dec. 2009 16 9.2 – 15.9 Domestic well 15.8 7.02 1726 162.7 80.8 90.0 40.3 0.02 0.2 0.27Z34 Dec. 2009 28 12.8 – 22.5 Agricultural irrigation 15.6 7.19 1985 175.4 102.9 1 05.0 0.4 0.02 0.3 0.28DW25 Dec. 2009 30 8.2 – 25.7 Domestic well 7.6 7.59 1412 119.0 67.8 111.0 0.5 0.02 0.3 0.24SUJL Dec. 2009 – River water 92.8 29.4 92.0 10.0 0.11 0.4 0.01ZC Dec. 2009 – Sewage water 81.0 28.9 168.0 14.2 0.02 0.4 0.11ZKE Jun. 2010 27 11.7– 22.8 Observation well 159.5 71.0 189.4 1.7 0.17 2.7 2.1ZKS Jun. 2010 27 11.2– 25.1 Observation well 103.4 35.5 151.5 2.1 0.13 4.0 1.4ZKW Jun. 2010 27 12.6 – 25.3 Observation well 227.7 101.6 80.5 7.7 0.04 0.7 0.24ZKN Jun. 2010 27 13.7– 24.1 Observation well 327.1 174.0 321.3 51.2 0.02 1.3 0.68

ID Samplingtime

Cd(μ g/L)

Pb(μ g/L)

Zn(μ g/L)

Cr6+

(μ g/L)Hg(μ g/L)

As(μ g/L)

Se(μ g/L)

NH4+ H2SiO3 CODMn Cl− SO42 − HCO3

− NO3− F− NO2

− I− Hardness TDS(g/L)

ZA Dec. 2008 0.9 2 11 b 4 0.1 1 0.2 0.17 20.1 0.52 55.9 46.9 506.0 2.4 0.63 0.004 0.10 492 0.56ZB Dec. 2008 1.2 3 2 b 4 0.1 1 0.2 0.03 20.1 0.76 62.2 89.8 315.0 0.2 1.11 0.004 0.06 312 0.47ZC Dec. 2008 2.0 2 87 b 4 b 0.1 b 1.0 b 0.2 47.9 13.41 153.0 120.0 631.0 b 0.2 0.77 0.014 0.09 321 0.89ZD Dec. 2008 1.0 2 22 32 b 0.1 1.9 b 0.2 0.25 6.97 53.9 52.8 274.0 b 0.2 0.86 0.040 0.02 224 0.39ZE Dec. 2008 2.6 3 29 b 4 b 0.1 10.6 b 0.2 141 40.31 328.0 380.0 656.0 b 0.2 0.49 0.013 0.10 838 1.62ZF Dec. 2008 2.7 4 18 b 4 0.1 1.1 0.2 0.02 28.6 1.47 140.0 277.0 777.0 126 0.43 0.480 0.01 932 1.39ZG Dec. 2008 2.2 2 41 b 4 0.1 1 0.2 0.02 20.9 0.68 115.0 155.0 698.0 260 0.41 0.004 0.02 829 1.28Z6 Dec. 2008 2.2 2 30 b 4 3.9 b 0.1 0.2 0.22 21.9 1.15 151.0 66.0 720.0 0.2 b 0.10 b 0.004 0.11 659 0.91Z6A Dec. 2008 4.0 13 30 5 0.6 b 0.1 b 0.1 b 0.02 29.8 202.8 58.3 521.1 1.3 0.33 0.005 0.20 634 0.78SW29 Dec. 2008 3.6 8 30 5 8.2 b 0.1 b 0.1 b 0.02 20.4 1.7 56.4 79.8 569.3 5.8 0.56 0.005 0.11 363.8 0.66SW49 Dec. 2008 4.4 16 37 5 b 0.1 b 0.1 7.2 b 0.02 20.1 1.8 246.7 126.0 660.2 2 94 0.46 0.044 0.01 1032 1.49SW59 Dec. 2008 b 0.1 15 30 5 3.7 b 0.1 b 0.1 b 0.02 11.2 1.8 126.9 300.0 297.2 0.4 0.92 0.005 0.19 511.9 0.84ZK1 Dec. 2008 4.0 2 5 5 1.7 b 0.1 b 0.1 b 0.02 31.2 1.5 104.2 73.5 314.9 0.1 0.45 0.005 0.01 142.1 0.53ZKE May 2009 1.1 4 5 b 4 b 0.1 4.4 1.4 0.45 26.6 2.3 164.0 147.0 691.0 b 0.2 1.33 0.082 0.02 399 1.00ZKW May 2009 0.4 3 3 7 0.2 3.2 1.4 0.33 24.4 3.5 51.2 143.0 179.0 15.9 1.16 1.980 b 0.01 155 0.46ZKS May 2009 0.6 4 2 7 b 0.1 23.9 0.9 0.99 24.7 3.9 122.0 137.0 478.0 24.4 1.83 0.730 0.02 230 0.82ZKN May 2009 1.0 4 3 b 4 b 0.1 3 0.8 0.02 27.7 7.2 976.0 149.0 599.0 b 0.2 b 0.02 0.012 1.02 1188 2.19ZF May 2009 0.7 3 1 b 4 b 0.1 0.9 1.2 0.13 23.3 1.6 138.0 331.0 644.0 110.0 2.25 0.013 b 0.01 680 1.26

(continued on next page )

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The wells were purged before sampling and groundwater was col-lected by pumping, only after constant values of electrical conductivity(EC)and redox potential (Eh)had been established. Most selected sam-pling points for groundwater were situated near the land ll. A freshsample tube was used for each piezometer to prevent cross contamina-tion. Water table depth, pH, turbidity, EC, Eh, dissolved oxygen (DO)and temperature were all measured in the eld. pH and temperaturewere measured at the time of sampling. All samples were lteredthrough a b 0.45 μ m membrane lter to separate the particulate fromthe dissolved fraction. In most cases alkalinity was determined on l-tered samples in the eld, by titration with H 2SO4 (0.22 N). Samplesfor major anion analysis were collected in polyethylene bottles,tightly capped and stored at 4 °C until analysis. Samples for cationanalysis were preserved in acid-washed polyethylene bottles, andacidi ed to pH b 2 with 6 N HNO3 . Major anions (Cl − , SO4

2 − , NO3− ,

precision ± 5 – 10%) were measured using ion chromatography (DX-

120), while major cations (Ca2+

, Na+

, K+

, Mg2+

, precision ± 5 – 10%)and minor/trace elements were analyzed using ICP-MS in theEnvironmental Chemistry Laboratory, School of EnvironmentalStudies, China University of Geosciences. The measured physico-chemical values of water samples from the study area can be seenin Table 1 .

Stable isotope values of δ18 O and δ2H were measured by mass spec-trometry, with a Finnigan MAT251 after on-line pyrolysis with a Ther-mo Finnigan TC/EA (Temperature Conversion Elemental Analyzer), intheState Key Laboratoryof Geological Processes andMineral Resources,China University of Sciences. The δ18 O and δ2H values were measuredrelative to international standards that were calibrated using V-SMOW(Vienna Standard Mean Ocean Water) and reported in conventional δ(‰ ) notation. The analytical precision for δ2H is ±1.5 ‰ and for δ18 O

is ±0.2 ‰ . The tested results are shown in Table 2 .

3. Results and discussion

3.1. Groundwater ow system

The land ll is located on the oodplain of the Jialu River. Alluvialsediments at the site range in size from coarse gravel to clay, althoughmedium-grained sand dominates the aquifers. Two alluvial aquifersare present within 0 – 52 m depth. The shallow aquifer consists of ne

Table 1 (continued )

ID Samplingtime

Cd(μ g/L)

Pb(μ g/L)

Zn(μ g/L)

Cr6+

(μ g/L)Hg(μ g/L)

As(μ g/L)

Se(μ g/L)

NH4+ H2 SiO3 CODMn Cl− SO42 − HCO3

− NO3− F− NO2

− I− Hardness TDS(g/L)

DW09 May 2009 0.6 3 1 b 4 b 0.1 1 1.8 0.14 19.3 1 187.0 515.0 771.0 487 4.19 0.680 0.01 234 1.99DW18 May 2009 1.2 5 1 14 b 0.1 7.6 1.3 0.21 13.4 3.2 249.0 225.0 424.0 101 2.59 1.790 b 0.01 608 1.11DW23 May 2009 0.8 3 1 b 4 0.2 1 1.6 0.04 21.1 1.4 81.5 279.0 604.0 85.5 2.50 b 0.004 b 0.01 501 1.11DW25 May 2009 0.7 6 4 5 0.2 1 0.9 0.06 19.3 1.3 178 140.0 548.0 b 0.2 1.63 b 0.004 0.08 305 0.94ZG10 May 2009 0.6 6 9 b 4 0.2 14.9 2.1 0.55 23.7 2.3 1159 203.0 5 21.1 b 0.2 b 0.02 0.008 1.81 2045 0.91ZG11 May 2009 0.7 5 6 22 b 0.1 4.8 1.4 75.6 14.6 46.7 615.0 249.0 995.0 b 0.2 2.30 b 0.004 1.59 725 2.36

ZG12 May 2009 0.7 3 4 5 0.1 b 1 1.6 b 0.02 32.2 11.4 140.0 242.0 357.0 43.6 1.83 0.005 b 0.01 457 1.27Z16 May 2009 0.7 4 1 4 0.2 b 1 2.0 0.11 17.3 1.0 244.0 445.0 342.0 156 1.58 0.005 b 0.01 705 1.31Z34 May 2009 0.7 5 3 5 b 0.1 b 1 0.9 0.08 20.2 2.0 200.0 222.0 484.0 34.6 1.91 1.570 0.01 475 1.12ZC May 2009 0.6 5 16 8 b 0.1 1.1 1.2 45.6 17.6 11.9 118.0 292.0 650.0 b 0.2 2.35 0.012 0.09 528 1.05SUJL May 2009 0.5 6 2 15 b 0.1 4.5 1.9 0.11 10.6 6 136.0 220.0 293.0 20.1 0.98 4.450 0.05 314 0.81SULG May 2009 0.5 4 3 7 0.1 8.1 1.9 0.32 b 2.0 7.2 131.0 207.0 227.0 b 0.2 1.48 0.043 0.13 224 0.71SUY May 2009 0.6 6 3 8 b 0.1 4.9 2 0.4 12.2 6 146.0 254.0 305.0 42.9 2.63 6.990 0.04 342 0.90SULD May 2009 0.7 5 6 18 b 0.1 7.1 2 16.4 15.1 7 155.0 184.0 526.0 41.2 2.66 0.006 0.30 423 0.96ZKE Dec. 2009 1.8 12 36 5 0.05 8.0 0.2 0.36 25.6 1.0 154.2 75.0 782.9 0.84 0.29 0.018 0.10 568.4 1.00ZKW Dec. 2009 0.6 3 10 5 0.01 3.6 0.2 b 0.02 30.1 1.3 110.6 131.0 713.3 85.0 0.21 0.077 0.02 792.6 1.02ZKS Dec. 2009 0.6 11 24 5 0.05 52.0 0.2 0.07 30.4 2.0 136.8 39.8 630.9 0.1 0.38 0.022 0.16 168.6 0.78ZKN Dec. 2009 8.2 38 65 5 0.03 6.4 0.7 0.04 25.3 1.0 1003.0 32.4 1578 3.7 0.38 0.026 0.01 1870 2.85ZB Dec. 2009 0.6 2 b 1 5 0.1 1.7 0.2 0.4 22.7 2.0 82.2 85.5 599.8 0.6 0.69 0.030 0.07 515.4 0.71ZG9 Dec. 2009 0.6 8 26 5 b 0.1 7.1 0.1 0.06 19.1 0.9 143.6 12.5 536.4 0.7 0.41 0.024 b 0.01 309.2 0.67Z06 Dec. 2009 0.7 2 3 5 b 0.1 0.2 0.4 0.02 22.7 1.2 48.9 52.0 549.2 195.0 0.39 0.016 0.01 628 0.83Z40 Dec. 2009 b 0.1 2 b 1 5 b 0.1 0.5 0.3 0.02 23.2 5.0 126.2 104.0 605.9 51.0 0.38 0.012 0.04 634 0.88Z16 Dec. 2009 0.8 2 20 5 0.04 11 0.2 1.39 20.9 9.6 38.6 54.5 429.0 1.7 0.46 0.118 b 0.01 368.3 0.48

ZF Dec. 2009 0.3 3 b

1 5 0.01 1.5 0.1 0.4 25.1 2.2 108.8 72.0 745.0 0.9 0.47 0.011 b

0.01 556.9 0.87DW09 Dec. 2009 0.5 2 1 5 0.04 0.8 0.5 0.02 20.3 0.4 110.6 177.0 719.4 70.6 0.30 0.106 0.01 741.1 1.09Z34 Dec. 2009 0.5 2 2 5 0.03 0.9 0.1 0.02 24.0 1.0 164.8 224.0 688.3 86.0 0.45 0.008 0.01 860.2 1.20DW25 Dec. 2009 0.7 3 30 5 0.01 2.4 0.1 0.02 19.4 0.1 68.4 149.0 612.0 93.5 0.06 0.003 0.06 578.5 0.92SUJL Dec. 2009 0.1 1 1 5 0.04 2.4 1.5 0.07 17.0 4.0 108.8 126.0 328.3 19.8 0.58 0.170 0.03 356.8 0.64ZC Dec. 2009 0.3 3 32 5 0.08 0.4 0.4 0.32 20.9 19.0 150.7 88.0 289.8 12.1 0.46 0.101 0.20 321.2 0.69ZKE Jun. 2010 0.8 2.1 7.4 b 4 0.01 14 – – 25.2 4.3 209.6 145.8 813.4 0.8 0.43 0.034 0.11 612.5 1.18ZKS Jun. 2010 0.1 1.1 4.3 b 4 0.08 24 – – 26.5 2.9 118.1 47.8 637.8 0.7 0.50 0.058 0.11 404.3 0.78ZKW Jun. 2010 0.6 3.7 1.0 b 4 0.02 0.9 – – 31.2 1.0 138.1 167.0 860.2 125.5 0.31 0.105 b 0.01 712.7 1.28ZKN Jun. 2010 3.1 14 9.6 b 4 0.03 1.9 – – 39.1 10.3 716.9 97.8 1358.0 10.8 0.27 0.030 0.68 1521 2.38

All in mg/L (if not speci ed otherwise) except pH.

Table 2Isotope composition for water samples around the Zhoukou land ll (sampled inDecember 2009).

Water sample Well depth (m) δ18 O (‰ ) δ2H (‰ )

ZKE 27 − 7.1 − 63ZKS 27 − 7.2 − 59ZKW 27 − 6.2 − 51ZKN 27 − 2.7 − 34ZB 50 − 6.6 − 57ZF 18 − 5.5 − 53Z16 30 − 3.7 − 36Z40 30 − 6.1 − 55Z06 15 − 6.8 − 60Z34 28 − 6.9 − 62DW09 19 − 6.0 − 55DW25 30 − 6.4 − 58ZG9 9 − 7.0 − 64ZK1 300 − 7.5 − 71SW49 14 − 6.7 − 59ZC Sewage water − 6.9 − 59SUJL Jialu River − 5.5 − 52

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sand and silty sand at a depth range of less than 30 m, and is a majorlayer for irrigation extraction in the local region. The deeper aquifer oc-curs at a depth range of 45 – 52m, and consists of ne sand. Low perme-ability silty clay of 11 m thickness exists above the shallow aquifer,which has a water table depth of 2 – 3m. TheinvestigatedZhoukou land-

ll with a thickness of 9.3m isembeddedin thissilty clay, whichis likelyto be an aquitard. The thickness of low permeable silty clay underlyingthe land ll is approximately 1.7m. Another silt and silty clay layer 20m

thick occurs between the shallow and deeper aquifers. The aquifers arewidely distributed across the study area, and their thickness becomesslightly thinner towards the south where the sediments become moreclay rich. In order to make comparison with groundwater hydro-chemistry at different depths, apart from the two shallow aquifers,onedeep well (ZK1, Fig. 1) reaching 300m in depth was also investigat-ed for hydrochemical analysis. The depth to groundwater in this well isabout 42 m below the ground surface, probably due to drawdowncaused by pumping in the deep layers.

The water table monitoring data shows that groundwater levels inthe study area varied from 2.3m to 8.2m depth. The hydraulic gradientof the shallow groundwater is approximately 0.0014 and locally nearthe land ll, higher than this. The local water table gradients belowand around the land ll are likely to differ from the regional gradient,as land lls usually have different hydraulic properties compared tothe surrounding area. A local water table mound has been observed atthe Zhoukou Land ll (water levels are higher than the surroundingareas). This could be due to higher in ltration capacity and preferentialrecharge at thesite, or pumping of groundwater in theareas outside theland ll. Groundwater levels in May 2009 were generally lower thanin December 2009 (after the wet season). According to the monitoringdata, the amplitude of the seasonal variation between wet and dryperiods ranges from 0.1 to 0.6 m.

Recharge to the alluvial aquifers consists of in ltration from precip-itation, rivers and channels, and leakage from irrigation return owsnear the land ll. Groundwater discharges to small streams, wetlands,and the Jialu River, and is subject to evapotranspiration. Groundwatersinks include discharge via evapotranspiration into the shallow soilzone, arti cial extraction, and out- ux of groundwater to surface waterin downstreamareas. With theexception of ood periods (where surfacewaters recharge the groundwater), groundwater discharges to the JialuRiver at all times. As Zhoukou city is located in the south of the YellowRiver alluvial fan, the saturated zone sediments are coarse-grained,which creates a potential pathway for rapid transport of contamination.These coarse-grained sediments are common throughout the HuaiheRiver Basin and are typically associated with groundwater pollution(Qu, 2010).

3.2. Stable isotope compositions

The δ18 O and δ2 H stable isotope composition of groundwater canbe analyzed in order to understand the recharge origins and circula-tion of the groundwater, and evaluate any possible evaporation(Table 2 ). Stable isotopic compositions of groundwater range from− 7.5 to − 2.7‰ for δ18 O and from − 71 to − 34‰ for δ2H. The meanδ2H and δ18 O values are − 56 and − 6.1‰ for groundwater and − 56

and −

6.2‰

for surface water, respectively.A plotof δ2H vs. δ18 O for surface and groundwater is shown in Fig. 3.The global meteoric water line (GMWL) corresponds to Craig (1961)and the local meteoric water line (LMWL, δ2H=7.3 δ18 O+ 2.1, n= 11fromJan. to Dec., r 2 =0.94) is based on δ18 O and δ2H values of the aver-age monthly rainfall monitoredbetween 1985 – 1992 at Zhengzhou sta-tion (34°43.2 ′N, 113°39.0 ′E) some 200 km NW of Zhoukou city. Therainfall data were taken from the IAEA network ( http://isohis.iaea.org). Most water samples are distributed along an evaporation line(Fig. 3), which is approximately parallel with the GMWL and LMWL. Awide variation in groundwater values from the study area on the rightside of the GMWL was observed. It is noted that this oxygen shiftcould be caused by elevation effects and/or evaporative processes tovarious extents prior to water reaching the saturated zone. The distribu-tions of δ2H and δ18 O in surface and groundwater samples are similar,andnear theLMWL, indicatingthatsurface andgroundwaterhave similarorigins (rainfall). Deep groundwater (from ZK; well depth of 300 m) haslower values of δ2H and δ18 O, indicating a colder climate duringrecharge. This is common in the deep aquifers of northern China, wheremuch of the groundwater was recharged during the late Pleistocene(Currell et al., 2012 ).

3.3. Inorganic macrocomponents in surface and groundwater

3.3.1. Physico-chemical characteristics of water samplesThe temperature of all water samples ranged from 7.6 to 22.7°C.The

groundwater was characterized by the pH values from 6.61 to 8.94.Groundwater was characterized by low turbidity (0.1 – 1.9, mean valueof 0.7) in the winter season and high turbidity (0.9 – 17.2, mean valueof 5.9) in the summer season. Sewage water and leachate samples re-corded a much higher turbidity of 79.2 – 90.1 with a mean value of 84,suggesting that they may have more in uence on local groundwaterduring the summer season. To some extent, turbidity can re ect thedegreeof contamination or leachate in uence in different water bodies.A wide range of electrical conductivity values (396 – approximately4000 μ S/cm) indicate presence of either contamination by leachate(expected to have high salinity), and/or evaporation of groundwaterprior to/during recharge. COD Mn , (expressed in oxygen mg/L), can beregarded as an aggregative indicator re ecting the extent of groundwater pollution by organic and reducing inorganic matter(MEP, 2002 ). The presence of relatively high COD Mn (range of 0.08 –

46.7mg/L, leachate= 40.3mg/L) in the collected groundwater samplesnear the land ll indicates the threat of severe contamination of

groundwater.Surface waters are generally Na-HCO 3 · SO4 or Ca-HCO3 · Cl type

(according to the classi cation of Stuyfzand, 1986) with a TDS concen-tration ranging from 0.64 to 0.96 g/L with a mean value of 0.81 g/L.Groundwater samples collected in December 2008 and December2009 were found to be Ca · Mg-HCO 3 (Cl) type, and in May 2009 and June2010werefoundto be slightly moreNa and Cl-rich. TDS concentra-tions of groundwater ranged from 0.46 to 2.85g/L,with a mean value of 1.11g/L. Land ll leachate (ZE) is Na· Ca · Mg-HCO 3 ·Cl·SO 4 type with1.62 g/L TDS, and sewage water (ZC and ZD) is Na-HCO 3 · Cl orNa · Mg · Ca-HCO 3 type with less than 1 g/L TDS. Hardness valuesranged from 142 to 2045 mg/L as CaCO 3 . Piper plots show that thewater type is closer to HCO 3-type, and hence more bene cial forusage, in the winter season compared to summer ( Fig. 4). Groundwater

salinity in the summer season (mean TDS concentration of 1.29 g/L)

Fig. 3. Stable isotopic composition of water samples in the study area (relative to VSMOW).Thelocalmeteoricwaterline (LMWL) represents themonthlyaveraged isotopiccompositionof long-term precipitation at the monitoring station Zhengzhou, being part of the Global

Network of Isotopes in Precipitation (GNIP).

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was higher than in the winter season (mean of 0.97 g/L), indicatinggreater amounts of leachate generated in the Zhoukou land ll insummer, which could be attributed to the rising water table and

subsequent mixing of leachate with groundwater water nearby. Someevaporative effects in summer may also be responsible for the higherTDS contents.

Fig. 4. Piper diagram showing the hydrochemical composition of water samples around the Zhoukou land ll aquifers. Groundwater samples show hydrochemical compositions close tothat of the surface water in the wet season.

Fig. 5. Variations of some chemical parameters with distances away from the Zhoukou land ll center. (a) NO 3− and Hardness; (b) TDS, F − , Cl− and SO4

2 − ; (c)CODMn and As; (d) Fe, Mn,

Ca and Mg. TDS, Hardness, Cl−

, CODMn , As, Fe, Mn, Ca and Mg have obvious decreasing trends with the increase in distance from the land ll center.

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Additionally, TDS, hardness and COD Mn of groundwater showed anobviousdecreasing trend with increasing distance from the land ll cen-ter ( Fig. 5a,b,c), and TDS showed a decreasing trend with increasingsampling depth ( Fig. 5.d). These indicate that the land ll contributessalinity, hardness and COD Mn to the adjacent groundwater. Theconcen-tration contours of several selected ions and elements in shallowgroundwater are shown in Fig. 6. It can be seen from these maps thattheconcentrationof Cl,NH 4 andMn (common characteristicsof leachate)

decline away from the land ll.

3.3.2. Major anionsA wide range of major ion compositions were observed,as shown in

the Piper plots ( Fig. 4). HCO3− and Cl− accounted for 18.8 – 78.5% (mean

value of 55.6%) and 11.8 – 71.9% (mean value of 26.3%), respectively, of total anions; similar to surface water (46.3% and 28.6%, respectively).Three groundwater samples collected from the well ZKN ( Table 1 ),exceeded the proposed drinking water quality standards for chloride.There is a de nite attenuation of chloride in groundwater down gradi-

ent of the land ll (Fig. 6a and a ′ ).

Fig. 6. Concentration contours for theCl − (a, a ′),NO3− (b, b ′),SO4

2 − (c, c ′),Mn(d, d ′ ), F− (e, e ′),andNH 4+ (f, f ′ ) of groundwater samplesfrom the shallowaquifer ( b 30m depth), collected in

May 2009 and Dec.2009, respectively. Units are in mg/L.

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Fluoride (F − ) concentrations in groundwater range from 0.02 to4.19 mg/L (mean value of 0.93 mg/L) and are from 0.46 to 2.66 mg/L (mean value of 1.3 mg/L) in the surface water samples. The F − concen-trations in 33%of groundwater samples exceeded thedrinking standardof 1 mg/L. Among the groundwater samples collected in December2008, only well ZB exceeded the maximum acceptable level, and therewere no wells that exceeded this standard in December 2009. However,the majority (83%) of groundwater samples collected in May2009 were

characterized by F−

concentrations greater than 1mg/L, indicating thatF− may be more susceptible to migrate in the aquifer during the sum-mer. In addition, ne-grained sediments and the relatively stagnantgroundwater ow in this region are more likely to result in the uorineenrichment ( Gong et al., 2010 ). Long-term ingestion of high- uoridegroundwater is the major reason for endemic uorosis in this area(Gong et al., 2010 ), a condition that affects human health by causingdamage to bones and joints (e.g. dental and skeletal ourosis) ( Cauleyand Murphy, 1995 ). The existence of this long-term condition in theregion indicates that uoride had a geochemical source present in theaquifer before the installation of Zhoukou land ll. Groundwater uorideconcentration showed a strong correlation with sulfate (SO 4

2 − ) con-centration ( Table 3 ), and relatively lower F − concentrations (0.49 –

0.86mg/L) were detected in the sewage water and the leachate (e.g. ZC,ZD, ZE). Although there is a big variation of F concentrations in May andDecember 2009, it also can be seen from the hydrochemical isolinesmap ( Fig. 6e,e′ ) that F − shows lower concentrations surrounding theland ll. In this study, high F groundwater ( N 1 mg/L) was observed ingroundwater with high pH values ( N 7.5); consistent with the ndingsof Cai (1999) , Gong et al. (2010) and Currell et al. (2011) . Therefore, if the land ll leachate isa signi cant sourceof alkalinity, this could indirect-ly lead to greater levels of mobilization of uoride from natural geologicalsources (e.g. Kim et al., 2012). Common F-bearing minerals (e.g., uorite,hornblende) were detected in Funiu Mountain area ( Gong et al., 2010 ),which is located in the southwest of Zhoukou city and about 200 kmaway from Zhoukou. Many hydrogeochemical investigationsshow F con-centrations of groundwater in Xuchang and Pingdingshan area, whichclose to the study area, range from 1 to 2 mg/L. However, the F contentsof groundwater are locally higher than 2 mg/L (even up to 4.19 mg/L atthe well DW09) in this study area. Apart from the geological backgroundsource, phosphatic fertilizers may be responsible for high concentrationsof uorite in groundwater ( Li et al., 2009; Shui and Zhang, 2009 ).

The majority of groundwater samples in this study were collectedfrom theshallow aquifer (less than 30m depth), andnitrate(NO 3

− ) con-centrations ranged from 0.1 to 487mg/L (mean value of 59 mg/L) with43% of these samples exceeding the Chinese standard of 20 mg/L NO 3

− .Surface water NO 3

− concentrations ranged from 0.1 to 42.9 mg/L

(mean value of 13.7 mg/L) and nitrate could not be detected in thesewage water nor the leachate (e.g. ZC, ZD, ZE). This indicates thatlocal agricultural activities (e.g. the use of N-fertilizers with precipita-tion in ltration or irrigation return ow) are the primary cause of theextensive groundwater nitrate contamination ( Zuo et al., 2007 ), as op-posed to any point source nitrate contamination from the land ll site.There is no signi cant negative correlationbetween NO 3

− andHCO3− ,in-

dicating little or no heterotrophic denitri cation, which would other-

wise continuously remove nitrate from the aquifer ( Mohamed et al.,2003 ). This is likely due to a lack of denitrifying bacteria, which usetheoxygenfromnitrate to generate energyfrom organiccarbon, releasingbicarbonate and nitrogen gas to the system as byproducts ( Babiker et al.,2004 ). In the present study, groundwater collected from wells used foragricultural irrigation and domestic use, has relatively high NO 3

− and F−

concentrations.Biochemical reactions can produce ammonium(NH 4

+ ) in groundwa-ter, indicating organic pollution providing electron donors ( Chapman,1992 ). Nitrate reductionpotentially occursat themixingzone of land llleachate and shallow nitrate containinggroundwater, with the presenceof ammonium and elevated nitrite at Z40, ZKW and Z16, indicating thepartial nitri cation of the ammonium plume due to in ltrating oxicwater. Nitrate reduction can cause disappearance of nitrate with depth(Postma et al., 1991 ), also potentially affecting the shallow monitoringwells (e.g., ZG10, ZG11) in this study. Groundwater with high nitrateconcentration (even up to 487 mg/L at DW09) was distributed in thedownstream area of the land ll, and groundwater with low nitrateconcentration ( b 0.2mg/L) was distributed near the land ll (e.g., ZG10,ZG11, ZKE, ZKN), indicating that denitri cation is likely a dominantredox process at thedownstream fringes of theplume. Thecontributionof aerobic and nitrate-reducing environments to natural attenuation of the major contaminants increases with the O 2 and NO3

− concentrationsin ambient groundwater,and theextentof mixingwith the leachate, re-spectively ( van Breukelen, 2003 ).

Based on the theory of Gibbs free energy, the oxidation sequence forthe major chemical components at neutral pH can be in the order: O 2 ,NO3

− , MnO2 , Fe(OH)3 , SO42 − and CO2 (van Breukelen, 2003; Wilson

et al., 2004 ). Most groundwater is characterized by dissolved oxygen(DO) of more than 1 mg/L, indicating there is the aerobic aquifer envi-ronment, and so aerobic oxidation of organic matter is probably still amajor process. In terms of changes in solute concentrations along thegroundwater owpath ( Fig. 6), Cl− , TDS, Fe, Mn, NH4

+ , and Ca2+ arehigh near the land ll site, decreasing as groundwater moves away fromthe land ll.

Additionally, while groundwater Cl − concentrations show an obvi-ous decreasing trend with increasing distance from the land ll center

Table 3Pearson correlation matrix for major ions and pH, EC and TDS of the groundwater around Zhoukou land ll.

EC pH Ca2+ Mg2+ Na+ K+ CODMn Cl− SO42 − HCO3− NO3− F− Hardness TDS

EC 1pH − 0.60 1Ca2+ 0.48 − 0.58 1Mg2+ 0.63 − 0.57 0.55 1Na+ 0.56 − 0.27 − 0.11 0.14 1K+ 0.37 − 0.45 0.19 0.54 0.10 1CODMn 0.36 − 0.20 − 0.16 − 0.12 0.82 − 0.01 1Cl− 0.65 − 0.42 0.39 0.74 0.46 0.51 0.06 1SO4

2 − 0.24 0.15 − 0.17 0.11 0.26 − 0.17 0.08 0.01 1HCO3

− 0.62 − 0.58 0.66 0.48 0.26 0.33 − 0.04 0.60 − 0.09 1NO3

− 0.23 0.00 0.18 0.06 − 0.09 − 0.10 − 0.14 − 0.09 0.54 0.12 1F− − 0.10 0.37 − 0.45 − 0.08 0.16 − 0.21 − 0.11 − 0.01 0.71 − 0.09 0.40 1Hardness 0 .78 − 0.72 0.60 0.61 0.60 0.42 0.56 0.55 0.03 0.54 0.01 − 0.33 1TDS 0.81 − 0.55 0.30 0.52 0.87 0.32 0.68 0.68 0.24 0.46 0.05 − 0.05 0.87 1

Correlation is signi cant at the 0.01 level.

Correlation is signi cant at the 0.05 level.

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(Fig. 5b and Fig. 6a, a′ ), groundwater NO 3− , SO4

2 − andF − concentrationsdo not follow this trend ( Fig. 6), indicating that pollution of NO 3

− , SO42 −

and F − may be due to local anthropogenic activities (e.g., fertilizer utili-zation, discharging domestic waste water into aquifer locally) and geo-logical background, rather than land ll leachate. NO 3

− , SO42 − and Cl−

concentrations in groundwater are characterized by a decreasingtrend with increasing sample depth, similarwith other ions ( Fig. 7); in-dicating that the combined solute load is high near the surface due tothe combination of anthropogenic in uences (agriculture and land llleachate). The highest concentrations were found at depths of lessthan 30 m, where the shallow aquifer is composed predominantly of

ne sand.

3.3.3. Major cationsNa and Ca are the most abundant cations in the groundwater with

concentrations of 30.2 – 872.0 mg/L (mean value of 175.6 mg/L) and13.7– 212.6 mg/L (mean value of 69.9 mg/L), respectively ( Table 1 ;Fig. 4). Ca and Mg in groundwater are characterized by an obviousdecreasing trend with increasing distance from the land ll center

(Fig. 5d). Additionally, the highest concentrations of Ca and Mgwere found within the shallow aquifer ( Fig. 7a). This indicates thathigher concentrations of these cations may be caused by leakage of land ll leachate into the groundwater. The high Mg concentrationsaround the land ll may result due to industrial waste like cosmetics,cement and textiles are being dumped into the land ll (McBean et al.,1995 ).

3.3.4. Fe and MnHigh contents of dissolved Fe and Mn can be produced by reducing

conditions generated in the land ll. Fe and Mn variations in water aremainly controlled by oxidation – reduction conditions and acid – alkalibalance ( Virkutyte and Sillanpää, 2006 ). Fe and Mn can be oxidized byanaerobicbacteria and under low pH conditions, concentrationswill in-

crease ( Pitt et al., 1999 ). In this study thelowest Fe (0.05mg/L) and Mn

(0.02 mg/L) concentrations ( Fig. 8b) were found in the groundwaterwith the highest pH (8.5 and 8.9, respectively). However, it would ap-pear that pH is not the only factor controlling the distribution of Fe

Fig. 7. Characteristics of some chemical parameters with well depth around the Zhoukou land ll. (a) Ca and Mg; (b) SO 42 − , Cl− and NO 3

− ; (c) Zn, Pb and As; (d) Fe, Mn and TDS.

Fig. 8. Scatter plots of Ca – Mg(a) and Fe – Mn (b)concentrations versus pH of groundwater

samples.

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and Mn, as an extremely high concentration of Fe (4.8mg/L) was foundin well Z16, where thepH wasslightly alkaline (7.7).The Mn concentra-tion in 76% of all groundwater samples exceeded the National Chinesedrinking water standard of 0.1mg/L. Fe and Mn have a negative correla-tion with SO 4

2 – concentrations in all groundwater samples ( Fig. 9a)which is not in agreement with Olías et al. (2004) , who determinedthat Mn has a strong correlation with sulfates resulting due to sul deoxidation. The Fe and Mn concentrations also showed an obvious de-creasing trend with increasing distance from the land ll center(Fig. 5d and Fig. 6) and with increasing sample depth ( Fig. 7d), indicat-ing higher concentrations of Fe and Mn caused by in ltration of reduc-ing leachate. In contact with the process of reducing leachate from theland ll, manganese and iron oxyhydroxide minerals in the subsoil willbe more soluble in a reducing environment than in thenatural oxidizingenvironment ( Deutsch, 1997 ). The low Eh conditions ( b − 50 mV de-tected in observation wells ZKE and ZKS in December 2009) can en-hance the dissolution of both the manganese and iron oxide and

hydroxide minerals beneath and downgradient of the land ll. Varia-tions of groundwater level in different seasons may change the redoxenvironment in the aquifer. Away from the land ll, the reconstructionof oxidizing conditions should effectively lower the dissolved concen-trations of Fe and Mn owing to the low solubility of their mineralsunder oxidizing, near-neutral pH conditions. Elevated levels of majorcations present in groundwater may partially result from cation ex-change reactions driven by the high levels of Fe, Mn, NH 4

+ near theland ll (Deutsch, 1997 ). Cation exchange with clays may be animportant process responsible for the high concentrations of NH 4

+ ,Fe, and Mn at land lls. As the concentrations of Fe, Mn, and NH 4

+

decrease away from the land ll, the typical major cations will bepreferred on the exchange sites. Additionally, there will be the effectson absorbed concentrations due to the low pH and Eh beneath the

land ll.

3.3.5. As, Se and Hg During the sampling campaigns, ZKS had high concentrations

(0.024 – 0.052 mg/L) of As, which were signi cantly higher than themaximum allowable concentration (0.01mg/L) according to the NationalChinese guidelines. ZG10, Z16 and ZKE recorded As concentrations of 14.9, 11.0 and 14.0 mg/L, respectively. If these values are excluded fromthedataset, themean concentrationof As in theremainingwater sampleswas 0.003 mg/L, which is below the maximum allowable concentration.Consistentwith the ndingsof Varsányi et al. (1991) , theFe concentrationin the groundwater strongly in uenced the distribution of As with thehighest concentrations of As (up to 0.052 mg/L) observed in Fe-rich(Fig. 10) and SO4

2 − -poor environments ( Fig. 9b). Additionally,groundwa-ter As concentrations were characterized by an obvious decreasing trendwith increasing distance from the land ll center ( Fig. 5c) and increasingsample depth( Fig. 7c). Ingeneral, Seand Hg werefound invery low con-centrations (up to 7.2 μ g/L and 0.2 μ g/L, respectively) withoutexceedanceof the NC standards.

3.3.6. Heavy metals of surface and groundwater Generally, land ll leachates contain only modest concentrations of

heavy metalsas themetalsaresubject to strong attenuation by sorptionand precipitation ( Christensen et al., 2001 ). Thezinc (Zn)concentrationrangedfrom below thedetected limit (BDL) to 0.065mg/L, with the ma- jority of samples recording a low concentration that did not exceed theWorld Health Organization (WHO) and National Chinese (NC) stan-dards.The presenceof zinc in thegroundwater indicatesthat theland llmay be receiving waste batteries and uorescent light bulbs and is acause for concern with regards to plant and aquatic life ( Al-Yaqoutand Hamoda, 2003 ). Aluminum (Al) levels ranged from b 0.02 to1.1 mg/L with 21% of groundwater samples exceeding the WHO andNC standards (0.2mg/L). It can be seen from Fig. 9b that low sulfate cor-responds to high Pb. In general, heavy metals of groundwater samplesare characterized by high concentrations in the monitoring wells as is

seen in the samples taken in December 2009 ( Fig. 11). Lead (Pb) con-centrations in the groundwater ranged from BDL ( b 0.002) to0.038 mg/L. Pb concentrations of N 0.01 mg/L were also found in somewells. Possible sources of lead contamination may be batteries,photographs, old lead-based paints and lead pipes disposed at the land-

ll, which are toxic to all forms of life at this level ( Al-Yaqout andHamoda, 2003 ). Acidity in leachate causes lead to be released from re-fuse, for example, the highest Pb concentration of 0.038mg/L occurredwith a pH value of 6.6 and was observed in well ZKN in December2009. Zn andPb concentrationsof groundwater samples were alsochar-acterized by an obvious decreasing trend with increasing sample depth(Fig. 7c). The highest concentrations of Pb were observed in the SO 4

2 − -poor environments ( Fig. 9b). Cd and Cr concentrations ranged fromBDL (b 0.001) – 0.008 mg/L and BDL (b 0.004) – 0.02 mg/L, respectively,

with the lower concentrations not exceeding the WHO and NC

Fig. 9. Scatterplotsof Fe – Mn(a) andAs – Pb (b)concentrationsversusSO 42 − concentrations of

groundwater samples.

Fig. 10. Scatter plots of Fe versus As contents, showing that the highest concentrations of As were observed in Fe-rich environments.

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standards. The high concentrations, however, indicate that therewill be some concerning human health risks caused by high in-take of Cd ( Ikem et al., 2002 ) and Cr (VI) (Wedeen and Qian,1991 ). Seasonal variations of the major heavy metals in ground-water ( Fig. 11) showed that these contaminants are charac-terized by higher concentrations in May 2009 and lower

concentrations in December 2008 and December 2009. General-ly, heavy metals do not seem to constitute a signi cant pollutionproblem at the land ll site, partly because concentrations in thesewage water and the leachate are often low, and partly becauseof strong attenuation by sorption and precipitation in the sedi-mentary aquifers ( Xie, 2011 ).

Fig. 11. Variation of contaminant concentrations in the selected monitoring wells during three sampling campaigns.

Fig. 12. Results of the principal component analysis (PCA) for major chemical components in groundwater samples surrounding the Zhoukou land ll site, showing the plots of variablesand loadings (a) and (a ′)—Principal component loading plot for the water samples collected in May 2009 and Dec.2009, respectively; (b) and (b ′)—component score plot for the water

samples collected in May 2009 and Dec.2009, respectively.

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3.4. Principal component analysis

To better distinguish the hydrogeochemical processes controllingthe groundwater hydrochemical composition, principal componentanalysis (PCA) (a common multivariate statistical method) was appliedto thegroundwater samples. Thetotal numberof components from PCAcan indicate potential sources of groundwater salinization and provideinsight into the interpretation of the data. In this study, the multi-

variables include Ca2+

, Mg2+

, Na+

, K+

, Fe, Mn, NH4

+

, SO4

2 −

, NO3

−

, Cl−

,F− , HCO3− , and TDS in groundwater samples (the dataset shown in

Table 1 ). Al, Pb, SiO2 and As variables were not included because theyare frequently below the quanti cation limit, and all analytical valuesless than dbl are replaced with zeros. The water samples collectedfrom the shallow aquifer (13 – 25 m depth) in different time periods(May 2009 and Dec. 2009) have been analyzed separately. The calcula-tion was carried out with SPSS 16.

After autoscaling, two signi cant components were identi ed torepresent 68.3% of total sample variance (43.3% for PC1 and 25.0% forPC2) in May 2009 ( Fig. 12a,b), and 73.7% of total samples variance(46.5% for PC1 and 27.2% for PC2) in December 2009 ( Fig. 12 a′ ,b′ ), re-spectively. The third component takes into account only 13.4% and9.1% of the total variance, respectively, and was not considered in thepresent analysis. Fig. 12 shows the loading plot and substantiates thatthe rst component is mainly related to Na, Ca, Mg, K, Cl and TDS inthe two sampling periods, whereas the second component is mainlyrelated to F, NO 3 , SO4 , and HCO3 in May.2009, and to Fe, Mn, and NH 4

in Dec. 2009.In this context, three groups of samples can be identi ed, seen in

Fig. 12a, a′ . In space variables, the PC1 factor takes high positive loadingfor the most chemical components in groundwater except for NO 3 , SO4 ,Fe, NH4 and F that show a negative loading. Group 1 can be described bymineralization of groundwater by water-rock interaction. The negativeloading for Group 2 and Group 3 that corresponds to high positiveloading for Group 1 may provide insight into the anthropogenicpollution, including leakage of land ll and fertilizer input through irri-gation in ltration along the groundwater ow paths surrounding theland ll.

The axis PC2opposes Fe, Mnand NH 4 toNO3 , SO4 andHCO3 . Howev-er, there are different loadings with opposite trend on these chemicalcomponents in different sampling periods, due to the groundwaterlevel uctuation and redoxenvironment variations.The opposite evolu-tion of Alkalinity (HCO3

− ) and Ca 2+ in PC2 in May 2009 indicates thattheincrease of Ca 2+ content in groundwater is likely related to gypsumdissolution and/or cation exchange and not carbonate weathering.There is also a notable decline in sulfate ( Fig. 6c,c′ ) near the land lldue to sulfate reduction. Compared with the ambient groundwater,the sulfate concentration in the land ll leachate has generally lowervalues, and may not maintain a degradation potential equal to iron re-duction. The elevated concentrations of dissolved Fe and Mn in ground-water near the land ll indicate these elements mainly mobilize underreducing conditions from land ll leachate.

Based on theanalysis of variable loadings on PC1 – 2 in Fig. 12a, a′ , thedifferent factors controlling on groundwater hydrochemical compo-nents can be proposed (see Fig. 12b, b′ ). The values of factor scoresare lower or higher than 0 depending on their relationship to theinten-sity of chemical processes that each factor represents ( Szynkiewiczet al., 2011 ). In summer season, the irrigation return ow and water-rock interaction make large contributions to groundwater quality to-wards the downgradient area (e.g., DW18, DW25, DW23, Z34), andthe leachate pollution plays thedominant role in controlling groundwa-ter quality surrounding the land ll site (e.g., ZKS, ZKE, ZKW). Amongthese water samples, DW18 with high score N +4 in PC2 is strongly as-sociated with nitrate pollution from agricultural practices. In the winterseason, the leachate pollution and the water-rock interaction give highscoresin PC2, havinggreat in uenceon groundwater quality around the

land ll site.

3.5. Contaminant transport model

In order to investigate the extent of migration of the land ll leachateinto the sand aquifer and the role of the clay/silt layers, a chloride trans-portmodel along a cross sectionwas constructed. Although groundwater

ow in the aquifer is essentially horizontal from the area north of theland ll to the area south of the land ll (Fig. 1), the zone of leachate con-tamination beneath the land ll has penetrated downward through the

entire upper sand aquifer thickness to a depth of 13–

25 m ( Han et al.,2013 ). This indicates that the vertical ow is also a primary in uence onthe distribution of contamination zone. The contaminant transportmodel was used to assess how land ll leachate may migrate into theuppersand unit, or whether the deeper sand unit couldbe contaminated.

Chloride ion was selected as the conservative tracer because itis generally not subject to undergo any physico-chemical reactions(e.g., adsorption, precipitation, andbiological degradation) in the aquifers(Christensen, 1992 ), andusuallyit is characterized by high concentrationsin leachate and leachate-contaminated groundwater near the land ll.Therefore, chloride can be used to study dispersion and dilution of a con-tami nant plume. Other conservative solutes are expected to behave in asimilar fashion and could be simulated, given data on in uent and back-ground concentrations. The investigation of organic contamination inshallow groundwater surrounding Zhoukou land ll showed that thereis no in ow of polluted groundwater into the river and sewage ditch,and the contaminated river and waste water from sewage ditch doesnot generally recharge the groundwater body ( Han et al., 2013 ).

The land ll body is the main pollution source for groundwater con-tamination in the study area. No other point sources or distributedsources of chloride ion were simulated in this study. The computercode MODFLOW-2005 ( Harbaugh, 2005 ) was used to simulate thegroundwater ow, and the computer code MT3DMS ( Zheng, 1999 ),was used to simulate the two-dimensional advective-dispersive trans-port of a conservative solute down gradient from the land ll site. The2D advection dispersion equation is written as ( Zheng, 1999 ):

∂ C ∂ t

¼ D x∂ 2C ∂ x2 þ D z

∂ 2C ∂ z 2

− v x∂ C ∂ x

− v z ∂C ∂ z ;

where C is dissolved concentration; t is time; D x , D z are dispersioncoef cient; v x, v z are velocity component in X and Z direction. The

3.5.1. Model geometry and parametersBased on theanalysis of groundwater hydrochemistry andhydraulic

dynamics in thestudy area, a reliable hydrogeologicalconceptual modelwas established, shown in Fig. 13. The N-S cross section was chosen toconstruct the contaminant vertical transport model ( Fig. 13a). Each of the silt/clay and sand units shown in Fig. 13a was discretized into 2 to4 model layers. The cross section was vertically discretized 19 layersin total, with a uniform horizontal 20 m grid cell size with thickness of 2– 3 m ( Fig. 13b). The top boundary of the cross section model receivesrecharge from precipitation. The bottom boundary of the model is re-

stricted by the depth of available investigation boreholes in this siteand is assumed no ow. Groundwater recharge in the study was esti-mated of ~ 36 mm/yr ( Qu, 2010), and the rate of leachate entry to theaquifer was estimated of 60 – 280 mm/yr ( Song et al., 2012 ). The ux

owing in/out in the lateral boundaries of the sand aquifer were es-timated according the geometry of the boundary and the regionalhydraulic gradient.

A representative value of the horizontal hydraulic gradients in theregional aquifer covering the land ll site is in the range of 0.0002 -0.00033 ( Qu, 2010 ). Based on the measured water level contours in De-cember 2009, the hydraulic gradient near the land ll is approximately0.0014. The values for hydraulic conductivity from 12 pumping testsin the ne- and medium grained sand are in the range of 11 – 12 m/d(Qu, 2010). The effective porosity value ranging 5 – 15% is anticipated

for the aquifer made up of ne sand in this site. Based on the data on

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hydraulic gradient (i), effective porosity ( n) and hydraulic conductivity(K ), values of the average groundwater seepage velocity ( v) were ob-tained as v= − Ki/n, in a range of 41 – 112m/yr.

No information is available about the hydraulic conductivity of the silty clay and silt units at this site. Thehydraulic conductivity valuesof the clay/silt layers are therefore estimated through sensitivity analy-sis by calculating water level distribution and travel time of leachate.Mechanical dispersion caused by velocity changes related toheterogeneity is generally negligible in porous media with a very lowhydraulicconductivitysuch as silt andclay. For thepurposeof modelingvertical migration of contaminants in the silt/clay layers, mechanicaldispersion is neglected and only advection and diffusion wereconsidered.

3.5.2. Model results and sensitivity analysisSensitivity analysis is conducted by running the model by varying

parameters of vertical hydraulic conductivity of the clay/silt layers andeffective porosity. The values of vertical hydraulic conductivity of theclay/silt layers were assumed to be in range of 0.0001 – 0.01m/d. Giventhe operation time of the land ll since 1998, the travel time throughthepathlinesstarting thebottom of the land llto the top ofupperaqui-fer serves as a useful response for model sensitivity analysis. The traveltime is calculated by the particle tracking software MODPATH ( Pollock,1994 ) and is a sensitive to the hydraulic conductivity and effective po-

rosity. The effects of uncertainty in vertical hydraulic conductivity andeffective porosity are shown in Table 4 . The sensitivity shows that vari-ation of vertical hydraulic conductivity in three orders of magnitude re-sults in about one order of magnitude change in the travel time.

Reducing the effective porosity to 1/3 of base value (0.15) also de-creased the travel time to about 1/3 of the base value.

The model is run 13years to represent the operation time of the land-ll since 1998 to 2010. Several model runs are made with longitudinal

dispersivity ( α L) as 1, 10 and 100 m. Vertical transverse dispersivity isassumed 1% of the longitudinal dispersivity. The simulated chlorideconcentration show signi cant different break-throughs in the upperaquifer beneath the right lower corner of the land ll (Fig. 14). Longitudi-naldispersivityof1m resulted ina maximum value of chloride concentra-tion of ~500 mg/L at the end of simulation, which is much lower thanchloride concentration observation. Although longitudinal dispersivityof 10 m and 100 m can produce chloride concentration of ~ 1000 g/L atthe end of 2010, the shape of the break-through is signi cantly different.For 100 m of dispersivity, the concentration reached about 760 mg/L only after 2yearsof simulation, which maynotbe anticipatedconsideringthe operation history of the land ll. The simulated values consideredreasonable for horizontal and vertical transverse dispersivities inthe sand aquifers are 10 m (longitudinal) and 0.1 m (vertical),respectively.

Fig. 13. (a) Model geometry for the transport model and (b) nite difference grid and boundaries for the ow model.

Table 4Average travel time under different Kv and effective porosity along the pathlines beneaththe land ll to the top of upper aquifer.

Kv (m/d) 0.0001 0.001 0.01 0.1

Travel time (yr) 110.7 23.4 22.0 13.4Effective porosity 0.05 0.1 0.15Travel time (yr) 7.3 14.7 22.0 Fig. 14. Concentration break-through in the upper aquifer beneath the right lower corner

of the land ll for different dispersivities.

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At the end of 13years, theestimated movement of theleachate front(concentration contour line of 10 mg/L) would be ~60 m to south as-suming a steady state ow. These estimates indicate the lateral migra-tion of the plume in the aquifer is limited. The simulated chlorideplume was constrained in the upper sand unit in the 13 year periodsince thebeginning of operation of the land ll (Fig. 15a). Another factorthat may affect the migration of plume beneath the land ll is the sea-sonal variation in water levels, which may result in variation in the ver-tical hydraulic gradient. Theaverage annual recharge used in the steadystate ow model (scenario 1) for transport simulation ( Fig. 15a) wastherefore redistributed into each season according to the ratio of sea-sonal precipitation and annual precipitation in each corresponding sea-son. The transient ow model (scenario 2) was applied to simulatecontaminant migration in the aquifer ( Fig. 15b) under these conditions.The model results indicate that the transient vertical hydraulic gradientmaybe the primary cause for themigration of contaminant beneath theland ll. Under the transient ow, the front of the chloride plume verti-cally reaches a deeper depth from the ground surface. The simulated~0.35 meter difference between springand summer water table causeda higher magnitude of vertical gradient over summer and autumn, andthereforeresulted in several metersof vertical migration of thecontam-inant plume.

4. Conclusions and implicated measures

The large number of uncontrolled land ll sites in China, such as thesite in Zhoukou is an increasing environmental and public health con-

cern, and has been identi ed as a major threat to groundwater quality.This study of the isotopic and hydrochemical composition of surfaceand groundwater around the Zhoukou municipal land ll has been con-ducted to better understand the extent of contamination from land llleachate, for a case where unlined and un-regulated lling occurred,acceptingtypical waste streams. Stable isotopes ( δ18 Oand δ2H) indicateelevation effects and/or evaporative processes prior to water reachingthe saturated zone, and show that surface and groundwater have simi-lar origins of recharge. Groundwater below and around the land ll sitehas been contaminated by leachate, clearly demonstrated in the pres-ence of reducing conditions and high concentrations of TDS, Cl − , NH4

+ ,FeandMn close to thesite,with obvious decreasing trendswith increas-ing distance from the land ll center, sample depth.

The analysis of NO 3− concentration distribution shows that agricul-

tural activities (e.g., the use of N-fertilizers) have also extensively

in uenced groundwater quality in the study area. Under low pH andEh conditions, the concentrations of dissolved metals (such as Fe andMn) in the groundwater increases. The data show the highest concen-trations of As were observed in Fe-rich environments, while F (probablypresent as a natural geological source) is more elevated in waters withhigh pH. Se, Zn, Cd, Cr and Hg are found in very low concentrationswithout exceedance of the WHO and NC standards, while As, Al andPb locally exceed these standards. In general, heavy metals do notseem to constitute a signi cant pollution concern at the land ll site,partly because the heavy metal concentrations in the sewage waterand leachate are often lower than the WHO and NC standards, and part-ly because of strong attenuation by sorption and precipitation in thesedimentary aquifers. TDS, Cl − , NH4

+ , Fe and Mn concentrations ingroundwater can be identi ed as the main indicators and make contri-bution to the groundwater pollution caused by land ll leachatepercolation.

A two-dimensional advective – dispersive transport model was estab-lished by using MODFLOW and MT3DMS to simulate the chlorideplume in the upper sand unit. Two scenarios (steady state and transient

ow model)were simulated anddiscussed. Chloride transport simulationshows that after 13 years operation of the land ll, the contaminationplume is still constrained in the upper aquifer. The seasonal dynamics of the hydraulic condition, such as recharge and water level temporal varia-tions should be investigated further to evaluate the migration of theplume beneath the land ll. Based on transient modeling, seasonal chang-es may be an important in uence on vertical hydraulic gradients andtherefore vertical ow of contaminants.

Generally, groundwater around the Zhoukou land ll is not suitablefor drinking water, and some remedial measures should be consideredto improve the groundwater quality around the municipal land ll, par-ticularlygiven the high levels of domestic and agricultural groundwaterusein this region. The land ll itself is non-engineered, without any bot-tom liner and leachate collection and treatment system. Therefore,leachate can easily nd a path into thesubsurface environment andcon-tinue to migrate in coming years. Some feasible options for improvinggroundwater quality areas follows: (i)Cover the land ll with an imper-meable clay in order to prevent rainwater from in ltrating the wastesite and reaching the land ll base. (ii) Extract and recycle the leachatefrom the land ll base so that less leachate will enter the underlyingaquifer. (iii) Treat andpurify thegroundwater around theland ll for ag-riculturalirrigation anddomesticuse inorder to eliminate thehigh NO 3

−

and F−

concentrations. (iv) Close the land ll. (v) Increasing the

Fig. 15. Simulated chlorideplume 13yearsafterthe implementationof land llin 1998 basedon (a) steadysate ow model and(b) transient ow model consideringseasonal variations ingroundwater level.

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evapotranspiration rate by planting vegetation cover overthe land llinorder to reduce leachate production.

Acknowledgments

This work has been carried out under the Exploratory Forefront Pro- ject (No. 2012QY007) for the Strategic Science Plan in the Institute of GeographicSciences and Natural ResourcesResearch,Chinese Academy

of Sciences, and was nancially supported as part of a groundwatersurvey project titled “Investigation and evaluation of typical contaminat-ed sites in Zhoukou region of Huaihe River Plain ” (no. 1212010634505).The authors are thankful to M.S. Xie Shiyong and Qu Zewei from ChinaUniversity of Geosciences for their assistance with eld work preparationandanalyzingwater samples in laboratory.Wealso thank theanonymousreviewers for their valuable comments and suggestions.

References

Abu-Rukah, Y., Al-Kofahi, O., 2001. The assessment of the effect of land ll leachate onground-water quality —a casestudy.El-Akader land ll site —north Jordan. J. AridEnvi-ron. 49, 615 – 630 .

Al-Yaqout, A.F., Hamoda, M.F., 2003. Evaluation of land ll leachate in arid climate —a casestudy. Environ. Int. 29, 593 – 600.

Arneth, J.-D., Milde, G., Kerndorff, H., Schleyer, R., 1989. Waste deposit in uences ongroundwater quality as a tool for waste type and site selection for nal storage qual-ity. In: Baccini, P. (Ed.), The land ll. , vol.20. Springer, Berlin, pp. 399 – 424 .

Babiker, I.S., Mohamed, A.A.M., Terao, H., Kato, K., Ohta, K., 2004. Assessment of ground-water contamination by nitrate leaching from intensive vegetable cultivation usinggeographical information system. Environ. Int. 29, 1009 – 1017.

Barrett, M.H., Hiscock, K.M., Pedley, S., Lerner, D.N., Tellam, J.H., French, M.J., 1999. Markerspecies for indentifying urban groundwater recharge sources: a review and casestudy in Nottingham, UK. Water Res. 33 (14), 3083 – 3097 .

Cai, H., 1999. Analysis of high F groundwater and its geological environment in Fuyangarea, Anhui, China. Groundwater 21 (3), 129 – 132 (in Chinese with English abstract).

Cauley, J., Murphy, P., 1995. Effects of uoridated drinking water on bone mass and frac-tures: the study of osteoporotic fractures. J. Bone Miner. Res. 10 (7), 1076 – 1086 .

Chapman, D., 1992. Water quality assessments. Aguide to the use of biota, sediments andwater in environmental monitoring. UNESCO/WHO/UNEP, Chapman & Hall, London371– 460 .

Christensen, J.B., Jensen, D.L., Gron, C., Filip, Z., Christensen, T.H., 1998. Characterization of the dissolved organic carbon in land ll leachate —polluted groundwater. Water Res.32, 125 – 135 .

Christensen, T.H., 1992. Attenuation of leachate pollutants in groundwater, land lling of waste. In: Christensen, T.H., Cossu, R., Stegmann, R. (Eds.), Leachate. Elsevier AppliedScience, London, pp. 441 – 469 .

Christensen, T.H., Kjeldsen, P., Bjerg, P.L., Jensen, D.L., Christensen, J.B., Baun, A., Albrechtsen,H.-J., Heron, G., 2001. Biogeochemistry of land ll leachate plumes. Appl. Geochem. 16,659– 718.

Craig, H., 1961. Standard for reporting concentration of deuterium and oxygen-18 innatural water. Science 133, 1833 – 1834.

Cui, Y.C., Fu, T., 1998. Harm of water pollution and organic pollutants in drinking waterresources in China. Urban Environ.Urban Ecol. 11 (3), 23 – 25 (in Chinese) .

Currell, M., Cartwright, I., Raveggi, M., Han, D.M., 2011. Controls on elevated uoride andarsenic concentrations in groundwater from the Yuncheng Basin, China. Appl.Geochem. 26, 540 – 552 .

Currell, M.J., Han, D.M., Chen, Z.Y., Cartwright, I., 2012. Sustainability of groundwaterusage in northern China: Dependence on palaeowaters andimpacts on water quality,quantity and ecosystem health. Hydrol. Process. 26, 4050 – 4066 .

Deutsch, W.J., 1997. Groundwater geochemistry: fundamentals and applications to contam-ination. Chapter 10 —practical applications: land lls. Lewis Publishers, Boca Raton, Fla.(USA) 153– 181.

Flyhammar, P., 1995. Leachate quality and environmental effects at active Swedish mu-nicipal land ll. In: Cossu, R., Christensen, H.T., Stegmann, R. (Eds.), Regulations, Envi-ronmental Impact and Aftercare. Proceedings Sardinia' 95, Fifth International Land llSymposium (Sardinia, Italy), vol. III. CISA (Computer and Information SystemsAbstracts), Environmental Sanitary Engineering Centre, Cagliari, pp. 549 – 557.

Foster, S.S.D., 1998. The Urban Environment —evaluation of hydrological changes andtheir consequences. Proceedings of an International Symposium on Isotope Tech-niques in the Study of Environmental Change, IAEA, Vienna, 1997, pp. 321 – 338.

Gong, J.S., Ye, N.J., Ge, W.Y., Li, J.H., 2010. The relationship between uorine in geologicalenvironment and endemic uorosis in Huaihe River basin. Geol. China 37 (3),633 – 639 (in Chinese with English abstract).

Han, D.M., Tong, X.X., Jin, M.G., Emily, H., Tong, C.S., Song, X.F., 2013. Evaluation of organiccontamination in urban groundwater surrounding a municipal land ll, Zhoukou,China. Environ. Monit. Assess. 185 (4), 3413 – 3444 .

Harbaugh, A.W., 2005. Mod ow-2005, the US geological survey modular ground-watermodel the ground-water ow process. US Geol. Surv. Tech. Methods 2005, 6 .

Ikem, A., Odueyungbo, S., Egiebor, N.O., Nyavor, K., 2002. Chemical quality of bottled wa-ters from three cities in eastern Alabama. Sci. Total Environ. 285, 165 – 175.

Kim, K., Ko, K.-S., Kim, Y., Lee, K.-S., 2012. Co-contamination of arsenic and uoride in thegroundwater of unconsolidated aquifers under reducing environments. Chemosphere87, 851 –856.

Lee, J.Y., Cheon, J.Y., Kwon, H.P., Yoon, H.S., Lee, S.S., Kim, J.H., Park, J.K., Kim, C.G., 2006.Attenuation of land ll leachate at two uncontrolled land lls. Environ. Geol. 51,581– 593 .

Li, D.L., Ren, X., Zhang, W.Y., Na, J., Ma, J.F., 2009. Causes and characteristics of the uorideion in shallow groundwater in Xuyi area of the Huaihe River Basin. J. Jiansu Polytech.Univ. 21 (1), 42 – 45 (In Chinese with English abstract) .

Li, J., 2013. China gears up to tackle tainted water. Nature 499, 14 – 15.Looser, M.O., Parriaux, A., Bensimon, M., 1999. Land ll underground pollution detection

and characterization using inorganic traces. Water Res. 33, 3609 – 3616 .McBean, E.A., Rovers, F.A., Farquhar, G.J., 1995. Solid waste land ll engineering and de-sign. Prentice Hal, New Jersey (NJ) 521 .

MEP (Ministry of Environmental Protection of the People's Republic of China), 2002.Water and wastewatermonitoring andanalysis methods, Fourth edition.Chinese En-vironmental Science Press 223 – 228 .

Mohamed, M.A., Terao, H., Suzuki,R., Babiker, I.S., Ohta, K., Kaori, K., 2003. Natural denitri-cation in the Kakamigahara groundwater basin, Gifu prefecture, central Japan. Sci.

Total Environ. 307 (1 – 3), 191 – 2001.Morris, B.L., Darling, W.G., Gooddy, D.C., Litvak, R.G., Neumann, I., Nemaltseva, E.J.,

Poddubnaia, I., 2006. Assessing the extent of induced leakage to an urban aquiferusing environmental tracers: an example from Bishkek, capital of Kyrgyzstan, CentralAsia. Hydrogeol. J. 14, 225 – 243 .

Niininen, M., Kalliokoski, P., Pärjälä, E., 1994. Effect of organic contaminants in land ll leach-ates on groundwater quality in Finland.Groundwater quality management.Proceedingsof the GQM 93 Conference held at Tallinn, September 1993. , 220. IAHS Publ.no .

Olías, M., Nieto, J.M., Sarmiento, A.M., Cerón, J.C., Cánovas,C.R., 2004. Seasonalwater qual-ity variations in a river affected by acid mine drainage: the Odiel River (South WestSpain). Sci. Total Environ. 333, 267 – 281 .

Osenbrück, K., Gläser, H.-R., Knöller, K., Weise, S.M., Möder, M., Wennrich, R., Schirmer, M.,Reinstorf, F., Busch, W., Strauch, G., 2007. Sources and transport of selected organicmicropollutants in urban groundwater underlying the city of Halle (Saale), Germany.Water Res. 41, 3259 – 3270.

Pitt, R., Clark, S., Field, R., 1999. Groundwater contamination potential from stormwaterin ltration practices. Urban Water J. 1, 217 – 236 .

Pollock, D.W., 1994. User's Guide for MODPATH/MODPATH-PLOT, Version 3: a particletracking post-processing package for MODFLOW, the U.S. Geological Survey nite-difference ground-water ow model: U.S. Geological Survey Open-File Report. 94 – 464(6 ch.) .

Postma, D., Boesen, C., Kristiansen, H., Larsen, F., 1991. Nitrate reduction in an uncon nedsandy aquifer – water chemistry, reduction processes, and geochemical modeling.Water Resour. Res. 27 (8), 2027 – 2045 .

Qu, Z.W., 2010. Hydrogeological conditions and evaluation of groundwater resourcesin Yingbei region of Zhoukou city. China Univ. Geosci. 12 – 13 (Master thesis (inChinese)) .

Rapti-Caputo, D., Vaccaro, C., 2006. Geochemical evidences of land ll leachate in ground-water. Eng. Geol. 85, 111 – 121 .

Saarela, J., 2003. Pilot investigations of surface parts of three closed land lls and factorsaffecting them. Environ Monit. Assess. 84, 183 – 192 .

Shui, Y., Zhang, Y.Z., 2009. Groundwater quality investigation in the typicalarea of HuaiheRiver Basin. Zhihuai. 12, 13 – 14.

Song, Q.L., Jin, M.G., Xie, S.Y., Wang, H.M., 2012. Modeling analysis of leakage and COD of leachate from the land ll in Zhoukou northern suburb. Geol. Sci. Technol. Inf. 31 (6),157– 160 .

Szynkiewicz, A., Witcher, J.C., Modelska, M., Borrok, D.M., Pratt, L.M., 2011. Anthro-pogenic sulfate loads in the Rio Grande, New Mexico (USA). Chem. Geol. 2011(283), 194 – 209.

Tong, C.S., 2012. The characterization of groundwater pollution and prevention measuresaround an old land ll: heavy metals and polycyclic aromatic hydrocarbons. (Ph.D.degree thesis) China University of Geosciences (Wuhan), Chapter 2 ‘Site description ’

32– 35 (in Chinese with English abstract) .van Breukelen, B.M., 2003. Natural attenuation of land ll leachate: a combined biogeo-

chemical process analysis and microbial ecology approach. (Doctoral thesis) VrijeUniversiteit, Amsterdam 105 – 112 .

Varsányi, I., Fodré, Z., Bartha, A., 1991. Arsenic in drinking water and mortality in theSouthern Great Plain, Hungary. Environ. Geochem. Health 13, 14 – 22.

Virkutyte, J., Sillanpää, M., 2006. Chemical evaluation of potable water in Eastern QinghaiProvince, China: human health aspects. Environ. Int. 32, 80 – 86.Wedeen, R.P., Qian, L.F., 1991. Cr induced kidney disease. Environ. Health Perspect. 92,

71– 74.Wilson, R.D., Thornton, S.F., Mackay, D.M., 2004. Challenges in monitoring the natural at-

tenuation of spatially variable plumes. Biodegradation 15, 359 – 369.Xie,S.Y., 2011. Soilcolumn leachingexperiments of Zhoukouland ll typicalcontaminants

and risk assessment of groundwater pollution. China Univ. Geosci. 26 – 30 (Masterthesis (in Chinese)).

Zhao, H., Liu, M., Li, R., 2010. Investigation and evaluation and study on the utilization of the groundwater resources of Huaihe Watershed the groundwater resources of Huaihe Watershed. Groundwater 32 (3), 44 – 47 (in Chinese) .

Zheng, C.M., 1999. MT3D, A modular three-dimensional transport model for simulationof advection, dispersion and chemical reactions of contaminants in groundwatersystems. Report o the U.S. Envirnmental Protection Agency, p. 170 .

Zuo, Z.J., Luo, W.J., Wang, X.K., Wang, W.F., Wang, C.H., Liu, Y.T., 2007. The evolutionalcharacteristics of shallow groundwater of Sha-Ying section in Huaihe River. J. Geol.Hazards Environ. Preserv. 18 (3), 7 – 71.

39D. Han et al. / Journal of Geochemical Exploration 136 (2014) 24 – 39