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Mercury distribution and mobility at the abandonedPuhipuhi mercury mine, Northland, New ZealandCM Gionfriddoa, JM Ogorekb, M Butcherc, DP Krabbenhoftb & JW Moreaua

a School of Earth Sciences, University of Melbourne, Parkville, Australiab Wisconsin Water Science Center, US Geological Survey, Middleton, WI, USAc Department of Conservation, Whangarei Area Office, Whangarei, New ZealandPublished online: 27 Jan 2015.

To cite this article: CM Gionfriddo, JM Ogorek, M Butcher, DP Krabbenhoft & JW Moreau (2015): Mercury distributionand mobility at the abandoned Puhipuhi mercury mine, Northland, New Zealand, New Zealand Journal of Geology andGeophysics, DOI: 10.1080/00288306.2014.979840

To link to this article: http://dx.doi.org/10.1080/00288306.2014.979840

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SHORT COMMUNICATION

Mercury distribution and mobility at the abandoned Puhipuhi mercury mine, Northland, New Zealand

CM Gionfriddoa, JM Ogorekb, M Butcherc, DP Krabbenhoftb and JW Moreaua*aSchool of Earth Sciences, University of Melbourne, Parkville, Australia; bWisconsin Water Science Center, US Geological Survey, Middleton, WI,USA; cDepartment of Conservation, Whangarei Area Office, Whangarei, New Zealand

(Received 23 September 2013; accepted 7 September 2014)

The United Nations 2013 Minamata Convention advocates for updated environmental assessments of potential point-source sites ofmercury contamination, including historic closed or abandoned mines. The Puhipuhi mercury mine (Northland), a historic andabandoned mine site, is located near one of the headwaters of the Wairoa River. In this study, total mercury levels in soils and sediments(37.8–1748 µg g−1), total and methylated mercury in waters (69.6–240 ng L−1 and 0.637–1.95 ng L−1, respectively), and elementalmercury in air (0.06–0.5 ng m−3) were measured to assess the probability and form of mercury release from the site to the surroundingnatural environment. Results showed that mercury concentrations at the site are elevated compared to regional backgrounds and furtherwork is necessary to determine how mercury may be transported from the site.

Keywords: cinnabar; contaminant mobility; historical mine sites; hydrothermal; mercury contamination; methylmercury; New Zealand;ore deposit; Puhipuhi; remediation

Introduction

The 2013 enactment of the United Nations EnvironmentalProgram (UNEP) Legally Binding Instrument on Mercury,known as the Minamata Convention, renews worldwide focuson assessment and remediation of mercury-contaminated sites,such as historic mining operations that used mercury amalgamsor processed mercury ore (UNEP 2013). The Puhipuhi area inNorthland is one of only three places in New Zealand wheremercury (Hg) has been mined as the primary target commodity.The site is located within the Waikiore Conservation Area, northof Whāngārei, and is currently managed by the NZ Departmentof Conservation (Fig. 1). The Puhipuhi Hg mine (Fig. 2), non-operational for more than 60 years, is currently being assessedas a point source for Hg pollution to the region’s soil andwaterways. The Hikurangi depression, a former swampreclaimed for dairy farmland, receives discharge from thewestern Puhipuhi upland via the Waiotu Stream (Ferrar 1925;Henderson 1944; Northland Regional Council 2002). Thenorthern Wairoa River, which is the region’s largest river, alsoreceives discharge from the Puhipuhi upland via the WaiotuStream (Henderson 1944; Northland Regional Council 2002).

The Hg ore deposits of the Northland province of NewZealand are localised in active and fossil hydrothermal systems,particularly in the Ngāwhā and Puhipuhi areas, respectively. AtPuhipuhi, Hg is mostly present as cinnabar (HgS), but also as animpurity in marcasite (≤1 wt%), mackinawite (≤1 wt%) or Ba-phosphates (≤0.5 wt%) (Craw et al. 2000). Mercury miningoccurred at Puhipuhi Mine between 1917 and 1945, and fewprecautions were taken during this period to mitigate the spread

of Hg to the environment. Between 1917 and 1945 a total of43 tonnes (t) of Hg were produced from Puhipuhi ore (grading0.17–1.0%), with company-specific yields given below whenavailable (Christie et al. 1995). Between 1917 and 1921 theproperty was mined by Whāngārei Cinnabar Company, Auck-land Cinnabar Mining Company and NZ Quicksilver MinesLimited (NZQM), producing 15.5 t of Hg from a total of 1558 tof ore (Ferrar 1925; Butcher 2010). During this time the tenor ofthe Puhipuhi ore was described to be similar to deposits workedin California, but much lower than ore from Almaden, Spain(Ferrar 1925). From mine-workings and boring operations, theore-body was estimated to contain about 35,000 t of Hg ore(Ferrar 1925).

In 1926, the Great British Mine took over and producedc. 0.71 t of Hg from 400 t of ore in one year (Henderson 1944).From 1928 to 1933, no further ore was discovered and workceased. In 1934 Mercury Mines (New Zealand) bought therights, and between 1935 and 1938 1.24 t of Hg was produced(Webb 1940). In 1939 a syndicate, New Zealand MercuryMines Ltd. (NZMM), invested in infrastructure includingnew processing facilities with crushers, ore bins, rotary kilns,a condensing plant, auxiliary buildings and workshops (Webb1941). Waikiore Stream was dammed to ensure a water supply;Fig. 2 shows the site during the peak of its operation underNZMM in 1944 (Henderson 1944; Butcher 2010). The minewas active from 1941 to 1945, and produced c. 15.1 t of Hg fromover 4810 t of ore (Henderson 1944; O’Brien 1946; Butcher2010). Adverse weather conditions, a high ratio of overburden toore, the patchy nature of the ore-body, difficulties in securing

*Corresponding author. Email: jmoreau@unimelb.edu.au

New Zealand Journal of Geology and Geophysics, 2015http://dx.doi.org/10.1080/00288306.2014.979840

© 2015 The Royal Society of New Zealand

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earth-moving equipment and, above all, the collapse in the priceof Hg resulted in low production of Hg at Puhipuhi (O’Brien1946). Despite considerable investment and prolonged effort,Hg mining at Puhipuhi ceased totally in 1945. The site hasremained undisturbed, with much of the WWII-era processingequipment still standing. The mine is situated at the north end oftheWaikiore Stream and includes an open-pit quarry, dam and theremains of a processing plant (Fig. 3).

At Puhipuhi the Hg extraction process began with truck-loads of ore from the quarry being off-loaded onto a concretechute which led from the road to the crushers (Site G). The orewas crushed and then lifted to the furnace to be roasted (Site G).Gaseous elemental mercury (Hg0) released during roastinginitially passed through five horizontal metal pipes set into aconcrete reinforced metal trough (5.4 m × 10 m) filled withwater (Site G) to condense Hg0 (Butcher 2010). The metal pipesextended to the pipe battery (Site E), which consisted of tenlow concrete troughs (1.2 m × 3.6 m and 0.7 m apart), withterracotta pipes set into cooling troughs of water to condenseHg0 (Butcher 2010). Four metal condensing towers (12 mheight and 3 m diameter, Site C) trapped gaseous Hg0 thatpassed through the pipe battery, and funnelled condensed Hg0

back through the cooling system via the pipe exchange (Site D):

a hollow concrete structure that directed flow from the fivepipes of the cooling system to two larger pipes that fed into thetower structure. After the condensing towers, two drum-shapedmetal tanks (3 m × 3 m, Site B), and the exhauster fan set withina timber structure (Site A) trapped Hg0 and vented gaseous by-products of the Hg extraction process (Butcher 2010).

Only a few published studies of Hg contamination atPuhipuhi exist; a study by Hoggins & Brooks (1973) surveyedthe Hg content of molluscs collected from an estuary of theWairua (Wairoa) River, Northland, which receives input fromthe Puhipuhi watershed. Their results indicated a dispersion ofHg from Puhipuhi up to 35 km from the cinnabar deposits. Totalmercury (HgT) levels in water dropped below normal back-ground (0.0001 ppm) within c. 8 km; however, HgT levels insediments exceeded natural background levels (>0.01 ppm) upto 35 km downstream of the mine deposits. No other publishedvalues exist for HgT at or around the mine site, and no valuesfor methylmercury (MeHg) have ever been reported for thesediments and waterways around Puhipuhi. Public concern overuse of road aggregate sourced from Puhipuhi quarries promptedinvestigation of Hg release from quarry rock (Craw et al. 2002),and lead to the conclusion that sediment carried in runoff tostreams poses a higher risk of Hg contamination than leachingof Hg through dissolution.

Processed (i.e. roasted) Hg ore, known as calcines, containmore soluble Hg minerals (i.e. montroydite, eglestonite, corder-oite and elemental Hg) compared to unroasted ore (waste rockfrom the quarry), and leaching of calcines can be a significantsource of HgT as has been observed at numerous Hg mines(Gray 2003; Navarro et al. 2005; Stetson et al. 2009; Gray et al.2010; Wiederhold et al. 2013; Yin et al. 2013). Plans of mineworkings at Puhipuhi show two tip locations. In 1921 under NZQuicksilver Mines Ltd., the ‘mullock’ (i.e. waste rock) tip waslocated just south of the present-day dam, between the WaikioreStream and the quarry (Fig. 3; Ferrar 1925). In 1945 under theNZ Mercury Mines Ltd., overburden was dumped at the end ofthe tramway just above the Waikiore Stream in what was once anatural gully but has now been altered by tram-loads of over-burden (Butcher 2010). It is assumed that the tips at Puhipuhiinclude both waste rock (from the quarry) and calcines;however, historical records from Puhipuhi do not explicitlystate the contents. Published studies have shown that minewaste (waste rock and calcines) from historical Hg mine sitescan have HgT concentrations ranging from 10 to 1500 µg g−1

(Rytuba 2003). Methylmercury concentrations can also besignificantly inflated in calcined mine wastes (Gray et al.2002a, 2003). The presence of calcines will increase mobilityof Hg in leachate.

This report contains the first data describing total Hg insoils and waters and MeHg in waters at the abandoned andvegetation-overgrown Puhipuhi Hg mine. These data are placedinto the context of the historical mining operations at the site toidentify likely ‘hotspots’ for species-specific Hg contamination.The results of this study provide information regarding the

Figure 1 Location of the Puhipuhi Mine in Northland, New Zealand.

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potential for release of toxic Hg species from the mine to thesurrounding environment.

Materials and methods

Samples for Hg analyses were collected following protocolsoutlined in Olson and DeWild (1999) with the followingvariations. High-density polyethylene (HDPE) and Teflon®(PTFE) bottles were used for sample collection. Sample bottleswere cleaned by being filled with 6 M hydrochloric acid (HCl)

and heated overnight in a hot water bath (c. 40 °C), then rinsedwith reagent-grade water and stored in double zip-type bags.Water samples for total mercury (HgT) and total methylatedmercury (MeHgT) analyses were collected at both the mine siteand along a transect of the Waikiore Stream (Fig. 3) in HDPEand PTFE bottles, and filtered the same day through 0.45 μmmembrane syringe filters. GPS coordinates are provided forwater sample locations in Table 1. Filtered samples were storedin clean HDPE or PTFE bottles and spiked with 1% v/v 11 Mreagent-grade HCl and stored out of direct light at 4 °C, as per

Figure 2 Historic map of the mercury mining operation and processing plant at Puhipuhi (Henderson 1944).

Mercury contamination in New Zealand 3

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Olson & DeWild (1999). The PTFE bottles were used to acquireduplicate samples from two sites suspected to be low in Hg.Filtered water samples for common anion analysis were storedin sterile 50 mL conical tubes and kept refrigerated. Anionconcentrations (F−, Cl−, Br−, NO3

−, SO42−) were measured in

filtered water using a Dionex DX-120 ion chromatograph withan IonPAC As14 column (4 × 250 mm) in the Department ofChemistry at the University of Melbourne. An Orion 250acombination electrode was used in the field to take pH andoxidation-reduction potential (ORP, mV) measurements. TheORP measurements have been corrected to a standard hydrogenelectrode (+249 mV at 20 °C), as outlined in the manual for an

Orion 96–96 combination electrode, and are reported as redoxpotential (Eh, mV).

Surface soil samples were collected as a grab sample from adepth of roughly 5–10 cm from various areas of the processingplant using a stainless steel spade. The final sample was acomposite of three subsamples collected from the sampling site.Samples were placed in sterile 50 mL conical polypropylenetubes and sealed. Soil samples were stored on ice until theycould be stored in the laboratory at –80 °C. Sediments (water-suspended soil and debris) were collected from standing waterthroughout the processing area in sterile 50 mL conical poly-propylene tubes. Excess water was removed prior to analysis by

Figure 3 Map of the Puhipuhi Mine site with sampling locations. Arrow is pointing north.

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centrifuging the sample. A moss sample was collected from thespigot of a metal cooling trough near the kiln house remains.For solid HgT analysis, c. 5 g (wet-weight) of soil, sediment ormoss sample was freeze-dried overnight on a Heto-Drywinnervacuum system. Accurate GPS coordinates could not bemeasured in the processing area due to heavy tree cover; aclearing near the pipe battery (site E) provides the relativelocation of the processing area to the Waikiore Stream.

All water and freeze-dried solid samples (soil, sedimentsand moss) were shipped for analysis to the Wisconsin MercuryResearch Lab (WMRL) of the Wisconsin Water Science Center(USGS; Middleton, Wisconsin). Filtered HgT analysis of waterwas determined using cold vapour atomic fluorescence spectro-metry (CVAFS) following US EPA Method 1631 (Olson &DeWild 1999). Filtered MeHgT analysis of water was deter-mined by distillation, gas chromatography separation andspeciated isotope dilution mass spectrometry using inductivelycoupled plasma mass spectroscopy (ICP-MS) following USGSMethod 01–445 and WMRL protocols (Olson & DeWild 1999).Solid samples were analysed for HgT by acid digestion andCVAFS following USGS Method 5A-8 and WRML protocols(Olson & DeWild 1999; Olund 2004).

Mercury vapour concentrations in air directly above soilsand in processing equipment were determined at several areasof the mine using an oxidation-based Hg collection method,NIOSH Method 6009 (Eller & Cassinelli 1994). Samples werecollected on Anasorb® sorbent tubes, Model C300, containingHopcalite resin. Flow rate and volume of air sampled wereregulated using a SKC PCXR4 Universal Air Sampling Pumpwith a fixed flow rate of 2 L min−1, recommended for shortsampling times (Takaya et al. 2006). All samples were acquiredat the soil horizon or within structures, and sheltered from thewind. Once sampled, the sorbent tubes were sealed withTeflon® tape and then sent to the ChemCentre in WesternAustralia to be analysed by cold vapour atomic absorptionspectrometry per NIOSH Method 6009 (Eller & Cassi-nelli 1994).

Results

Water chemistry and Hg speciation data are shown for theWaikiore Stream and processing area at Puhipuhi Mine (Tables 1and 2). Around the processing area, the pH of standing waterwas acidic to neutral, ranging from 2.7 to 7.4; water trapped inthe cooling troughs had pH values within the range 5.2–6.4,similar to the background pH of 4.0–5.5 from the surroundingforested area (Craw et al. 2002). Concentrations of filtered HgTwere highest in the portion of stream below the dam (219–240 ng L−1) in samples PP1 and PP5, whereas above the damHgT was measured at c. 118 ng L−1. The lowest concentrationof HgT was found in the creek between the quarry and the damat 69.6 ng L−1. Filtered MeHgT concentrations were highestcloser to the dam (1.41–1.95 ng L−1). Further downstream,MeHgT concentrations were 0.637–0.823 ng L−1, similar tovalues measured above the dam (0.706–0.823 ng L−1).T

able

1Water

chem

istrydata

from

WaikioreStream

anddam

atPuh

ipuh

iMine,

includ

ingfiltered(<0.45

μm)mercury

(Hg T,MeH

g T)concentrations.

Sam

ple

site

Area

Location/

GPS

Coo

rdinates

pHEh

(mV)

Cl−

(mgL−1)

SO42−

(mgL−1)

NO3−

(mgL−1)

F−

(mgL−1)

Filtered

MeH

g T(ngL−1)

Filtered

Hg T

(ngL−1)

%Hg T

asMeH

g T

PP1

Waikiore

Stream

35°27′38″S,17

4°15

′8″E

5.1

364

6.54

7.45

<0.00

5n.d.

1.41

240

0.6

PP4

Dam

35°27′38″S,17

4°15

′3″E

7.1

245

7.90

1.73

0.24

n.d.

1.95

118

1.6

PP5

Waikiore

Stream

35°27′39″S,17

4°15

′10″E

5.1

271

7.80

9.08

<0.00

5n.d.

0.63

721

90.3

PP6

Abo

vedam

35°27′40″S,17

4°15

′3″E

7.2

320

9.19

20.8

<0.00

50.20

0.70

669

.61.0

PP5T

1Waikiore

Stream

35°27′39″S,17

4°15

′10″E

5.1

271

8.03

10.2

<0.00

5n.d.

0.82

3–

–

PP6T

1Abo

vedam

35°27′40″S,17

4°15

′3″E

7.2

320

9.55

20.4

n.d.

0.20

0.82

3–

–1PP5T

andPP6T

arereplicatesamples

foranionandmethy

lmercury

analyses,takenusingTeflon

bottles.

n.d.,no

tdetected.

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The HgT content of soils and sediments varied across theprocessing area at Puhipuhi (Table 2). The highest concentra-tions were found at Site D, the remains of the pipe exchange,and at Site A, the exhauster/fan (529–1748 μg g−1 and 37.8–1486 μg g−1, respectively). Sediments from the pipe battery,Site E, and the water holding tanks, Site F, had the lowestconcentrations of HgT (39.4 μg g−1 and 123 μg g−1, respect-ively). Sediments sampled from one of the condensing towers,Site C, had much higher HgT than soils collected from nearthe tower (353 μg g−1 versus 35.3–50.5 μg g−1). Very lowconcentrations of gaseous elemental mercury (Hg0) weredetected across the processing area (0.06–0.5 ng m−3). A back-ground air sample taken at Puhipuhi Rd near the entrance to themine site was below the detection limit (<0.01 µg Hg on thesorbent tube).

Discussion

Background HgT levels are typically of the order 0.04–74 ng L−1

for most Northern Hemisphere lakes and 1–7 ng L−1 for mostnorthern rivers and streams (USEPA 2001). Background MeHgTlevels in surficial waters range from 0.04 to 0.8 ng L−1 (Krab-benhoft et al. 2007). For New Zealand, a series of legislativestandards establish limits for HgT and MeHgT in drinking water,fish, and aquatic ecosystems based on these background con-centrations (Chrystall & Rumsby 2009). Measurements of thedistribution and abundance of HgTandMeHgTat the Puhipuhi Hgmine provide important data for comparison of Hg speciation andconcentration in local and regional waterways. Elevated filteredHgT levels in Waikiore Stream at Puhipuhi (c. 70–240 ng L−1),and the fractionation of HgT as MeHgT, are within the range ofpreviously published values for contaminated mine sites; seeTable 3 (Ganguli 2000; Rytuba 2000; Scudder et al. 2009;Kocman et al. 2010). Filtered water concentrations of HgT at

Puhipuhi are below New Zealand’s 0.007 mg L−1 (7000 ng L−1)maximum acceptable value (MAV) for inorganic Hg in drinkingwater (Ministry of Health 2013). However, they are above theANZECC and ARMCANZ (2000) water quality guideline triggervalue of 60 ng L−1 Hg for the protection of aquatic ecosystems.This value is lower than both the chronic continuous concentra-tion (CCC) for long-term exposure to aquatic organisms(770 ng L−1) and the chronic maximum concentration (CMC)for short-term exposures set by the US EPA (1400 ng L−1)(USEPA 2013).

Values for HgT upstream (PP6) of the tips are lower thanthose observed below the tips, indicating probable leaching ofHg from mine waste. Craw et al. (2000) estimated fromleaching experiments that 10–100 ppb of Hg could be releasedin low-pH (1–3) conditions from Puhipuhi quarry rock, andconcentrations of up to 100 ng L–1 HgT would be expected inwaters near mineralised deposits. Concentrations of HgT exceed100 ng L−1 in waters adjacent to both tips (PP1, PP4, PP5).Despite increased concentrations of HgT, MeHgT values aboveand below the dam are roughly the same. A low fraction (<1%)of HgT as MeHgT occurs in Puhipuhi stream waters, which istypical for streams in mined areas (versus un-mined areas)(Scudder et al. 2009). The dammed portion of the stream hadthe highest MeHgT concentrations (1.95 ng L−1) and highestfraction of HgT as MeHgT (1.6%), and has the greatest capacityto methylate Hg of the areas sampled, likely correlating tofactors such as dissolved organic matter, pH, chloride andsulphate (Compeau & Bartha 1985; Gilmour et al. 1998; Benoit& Gilmour 1999; Ullrich et al. 2001; Benoit et al. 2003). Theresults presented here are a preliminary snapshot; MeHgTmeasurements must be taken routinely for water and sedi-ments in order to understand the bioavailability of Hg from

Table 2 Total solid-phase mercury (HgT) content of soils and sediments, and gaseous elemental mercury (Hg0) from processing area, including pH ofstanding water when sediment samples were collected. Processing area is located at 35°27′43″S, 174°15′10″E.

Sample Area Description pH Hg0(ng m−3) Solid HgT(µg g−1)

PP003 Site A Soil, 5 cm under pipe near exhaust fan – – 1486PP009 Site A Soil, 5 cm depth, adjacent to exhaust fan – – 37.8PA1 Site A Gas, 10 min sample time, soil adjacent to exhaust pipe – 0.50 –PA2 Site A Gas, 30 min sample time, in exhaust fan – 0.33 –PA3 Site A Gas, 64 min sample time, in exhaust tower – 0.16 –PB1 Site B Gas, 49 min sample time, in tank on farthest right – 0.20 –PP005 Site C Soil, 10 cm depth, in front of second condensing tower – – 35.3PP006 Site C Soil, 8–10 cm depth, between condensing towers and pipe battery – – 50.5PP007 Site C Sediment from second condensing tower 4.3–4.6 – 353PC1 Site C Gas, 77 min sample time, in third condensing tower – 0.06 –PP001 Site D Ash-like sediment from inside concrete structure of the pipe exchange 2.8–3.1 – 774PP002 Site D Wet sediment from inside concrete structure of the pipe exchange 2.8–3.1 – 1748PP023 Site D Sediment from tiered pool of pipe exchange 2.8–3.1 – 529PP008 Site E Sediment from concrete cooling trough of pipe battery 5.2–6.4 – 39.4PP011 Site E Moss, sampled below spigot from concrete trough – – 9.1PP004 Site F Sediment from water holding tank 2.7 – 123

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Table 3 Total mercury (HgT), methylmercury (MeHgT) and gaseous elemental mercury (Hg0) of sediments, soils, water and vegetation of PuhipuhiMine, published values for other Hg mine sites, regional baselines and regulatory values.

Filtered Water Sediment/mine waste/ soil Air

HgT MeHgT HgT MeHgT Hg0

Location (ng L−1) (ng L−1) (µg g−1) (ng g−1) (ng m−3)

Puhipuhi MineWaikiore Stream at mine 69.6–240 0.61–1.02 – – –Downstream from mines, Wairua River1 100–1000 – <100 – –Soils and sediments (processing area) – – 35–1748 – –Puhipuhi mineralised rocks (Hg mine quarry)2 – – 1–100 – –Puhipuhi soils (C-horizon, >1 km from quarry)2 – – 0.01–100 – –Ambient air (processing area) – – – – 0.06–0.50Hg mines, Humboldt River Basin, Nevada, USA3

Downstream from mines 2.1–2000 <0.02–0.92 0.17–170 0.12–0.95 –Mine waste calcine – – 14–14,000 <0.05–96 –Coast Ranges Hg mines, California, USA4,5,6

Downstream from mines (San Carlos Creek) 1.63–140 0.017–1.2 91.3–319 – –Mine waste calcine and soil – – 380–6530 – –Processing area – – – – 3000–188,000Hg mines, Kuskokwin River, Alaska, USA7

Downstream from mines <50 <1.0 0.9–5500 0.05–31.0 –Mine-waste calcine and soil – – 3.5–46,000 0.2–41 –Hg mine, Palawan, Philippines8

Downstream from mines 8–13 – 3.7–15 0.28–3.9 –Mine waste calcine 100–30,000 – 28–660 0.13–3.2 –Hg mines, Monte Amiata, Italy9,10

Downstream from mines – – 0.26–14 0.2–8.7 –Mine waste calcine and soil – – 0.64–1500 0.34–45 –Buildings containing condensers and furnaces – – – – 5000–10000Idrija Hg mine, Slovenia11

Downstream from mines 0.25–143 0.12–0.27 0.77–4121 0.01–11 –Mine waste calcine and soil – – 0.3–19900 6.5–14 –Hg mines, Guizhou Province, China12

Mine waste calcine and sediments – – 5.7–4400 0.17–20 –Moss – – 1.0–95 0.21–20 –Comparative Standards13

NZ drinking water standard 7000 – – – –US EPA, acute aquatic life water standard 2400 – – – –US EPA, chronic aquatic life water standard 770 – – – –USEPA, regional screening level for soil – – 23 – –NZ MFE-ambient air quality guidelines, Hg (annual) – – – – 330NZ occupational exposure limits for inorganic Hg (8 hr) – – – – 25,000Background/baseline valuesFreshwater lakes14,15 0.04–74 0.003–6 – – –Freshwater rivers and streams14,15 1–7 0.078–0.55 – – –Global background for uncultivated soils16 – – 0.045–0.16 – –Auckland soils17 – – <0.03–0.45 – –Ngāwhā geothermal field (Northland) soils18 – – 0.017–0.35 – –South Island unweathered metagreywacke2 – – 0.1–0.25 – –South Island metamorphosed basalts2 – – 0.04–0.14 – –Typical Hg content in New Zealand ambient air13 – – – – 0.42–3.11Hoggins & Brooks (1973), 2Craw et al. (2002), 3Gray et al. (2002b), 4Rytuba (2000), 5Ganguli (2000), 6Degraff et al. (2007), 7Gray et al. (2000), 8Gray et al. (2003),9Rimondi et al. (2012), 10Vaselli et al. (2012), 11Kocman et al. (2010), 12Qiu et al. (2005), 13Chrystall & Rumsby (2009), 14USEPA (2001), 15Krabbenhoft et al. (2007),16Siegel (2002), 17ARC (2001), 18Davey & van Moort (1986).

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Puhipuhi to methylating bacteria, and the potential for MeHgTaccumulation in local aquatic or avian fauna and flora.

The global average of HgT in uncultivated soils falls withinthe range 0.045–0.16 μg g−1 (Siegel 2002). Natural concentra-tions of HgT in soil can depend greatly on the parent rock andmineral material which the soil was derived from. In NewZealand, volcanic soils and soils in geothermal areas (e.g.Auckland and the Ngāwhā geothermal field) have slightlyinflated concentrations of HgT compared to global values(<0.03–0.45 μg g−1 and 0.017–0.35 μg g−1 respectively)(Davey & van Moort 1986; ARC 2001). Craw et al. (2002)used rocks from the South Island that were comparable to thosefound in the Puhipuhi region to establish regional backgroundsof 0.1–0.25 ppm HgT for soils from unweathered metagrey-wacke and 0.04–0.14 ppm from clay-altered greywacke andmetamorphosed basalts. In the same study, mineralised rocksfrom Puhipuhi had HgT concentrations 1–100 ppm aboveregional background and the HgT content of Puhipuhi regionsubsoils (C-horizon) were highly variable, ranging from 0.1–100 ppm (Craw et al. 2002).

Soil and sediment HgT concentrations at Puhipuhi Mine areinflated compared to regional baselines; they are up to 2–3orders of magnitude above the global average and similar tovalues reported for mine surface soils (100–1000 μg g−1;Table 3). At the processing site, elevated concentrations ofHgT at Sites D and E (pipe exchange and pipe battery/coolingtroughs, respectively) are attributed to deposition of elementalmercury (Hg0) released during heating of the ore and by coolingand trapping of the Hg vapour. Elevated HgT concentrations insoils near condensing towers and the exhaust fan (Sites C andA, respectively) are most likely from emission deposition ofcondenser soot (fine-grained particulates of cinnabar that travelwith Hg vapour) and Hg0 that escaped condensation andcollection at Site D. Elevated concentrations of HgT areobserved away from towers and structures (Site C and Site F),which could be from physical placement of condenser soot,emission deposition of vapours, mechanical dispersion oraqueous dissolution of Hg0 and cinnabar in surface waters;these processes have been observed at other historic Hg minesites (Navarro et al. 2005; DeGraff et al. 2007).

Elemental Hg levels measured in the air within structuresand directly above soils in the processing plant area (0.06–0.50 ng m−3) are similar to typical ambient air in New Zealand(0.42–3.1 ng m−3), and well below the New Zealand ambientair quality guideline limit of 0.33 μg m−3 Hg0 (Chrystall &Rumsby 2009). These concentrations are much lower thanvalues measured at other historic Hg mines (>300 ng m−3)(Navarro et al. 2005; DeGraff et al. 2007; Gosar & Teršič 2012).The concentration of gaseous Hg at Site A (0.16–0.50 ng m−3)is several orders of magnitude lower than expected for soilthat contains upwards of 1000 ppm HgT. Considering thestandard Gibbs energy (ΔG°) of Hg0 vaporisation (Hg0(l) ⇔Hg0(g)) as 31.85 kJ mol−1, the Keq of the reaction at 15°C

(ΔG° = –RT ln Keq, R = 8.314 J mol−1 K−1) is 1.67� 10−6

(Linstrom & Mallard 2001). For a range of Hg0(l) of 30–1400 ppm, the expected equilibrium Hg0(g) would be c. 50–2300 ng m−3, similar to values measured at other mine sites(refer to Table 3). Sampling conditions at Puhipuhi do notprovide equilibrium measurements of Hg0(g), but a snapshot ofthe flux of Hg0(g) from soils and structures. Either the samplingis inefficient in accurately capturing the flux of Hg0 from thesoil, or the low concentration of Hg0(g) measured at Site Asuggests that the majority of HgT measured in the topsoil is notHg0(l) but matrix-bound Hg species. Precipitation of Hg withchloride, phosphate, carbonate and hydroxides from leachate,and adsorption of Hg species (e.g. Hg2+, Hg2

2+) by clayminerals, oxides and organic matter, are key processes in theretention of Hg in soils (McLean & Bledsoe 1992; Steinnes2013). The mobility of Hg species in soils is highly dependenton soil pH and redox conditions. In the acidic soils of Puhipuhiregion (pH 4–5.5), oxidised Hg species likely complex withchloride forming HgCl2

0(aq) or adsorb to organic matter,

inhibiting volatilisation of Hg0 from the soil matrix (McLean& Bledsoe 1992; Scudder et al. 2009; Steinnes 2013). Mercuryspecies originating from the host rock/ore could be present ascinnabar (HgS) which has low solubility at low pH, or as aminor impurity (< 1 wt%) in iron sulphide and bariumphosphate minerals which are highly soluble in low pHconditions (Craw et al. 2000).

Degassing of Hg from soils can fluctuate daily andseasonally, and continuous gaseous Hg monitoring (e.g. Tek-ran® 2537 automated ambient air analyser) can be expensiveand time consuming. Mosses are exposed to heavy metalsthrough wet and dry deposition (e.g. rain water or absorbed todust) rather than uptake from soil, and are indicators of Hgmobilisation in air at contaminated sites (Yeaple 1972). Enrichedconcentrations of HgT (>1 ppm) are typical for moss from Hg-contaminated sites (geothermal plants, chlor-alkali works andabandoned Hg mines; Yeaple 1972; Lodenius 1981; Baldi 1988;Qiu et al. 2005). Mercury was detectable in moss collected froma cooling trough at Site E (9.1 µg g−1) and further sampling ofmoss across the Puhipuhi site could be used to measure thedispersion of Hg in air. Mercury adsorbed to dust, rather thanvolatilisation of Hg0 from soil, may be a significant factor in theremobilisation and dispersion of Hg across the Puhipuhiprocessing site.

Lower pH waters (pH < 4.5) at the site (including standingwater in structures) are indicative of localised acid minedrainage formation (i.e. oxidative dissolution of metal sul-phides), and these waters may assist in the leaching of Hg fromsurficial soils during rainfall runoff (Gray et al. 2002b).However, erosion of soils and carry-off of soil during rainfallare likely to be greater factors in controlling the mobilisation ofHg across the site and to the Waikiore Stream. Further work,including geochemical analysis of surface soils (upper 10 cm)as well as sub-soils (C-horizon) and pH-controlled leaching

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experiments, would enhance our understanding of the speciationstate of Hg in Puhipuhi soils and how they may be mobilised.

Conclusions

At Hg mine sites, the three main pathways for release of Hg are:(1) gaseous elemental Hg from processing equipment and soils;(2) leaching of soluble Hg species from processing area soilsand mine waste; and (3) soil erosion and runoff of insoluble Hgspecies from the processing area and unroasted ore from minewaste. This study provides the first attempt at quantifying theabove pathways at the Puhipuhi mine and insights into futurework. The abandoned mine site at Puhipuhi follows similarcontamination patterns surveyed at Hg mine sites in the WesternUnited States (DeGraff et al. 2007). High HgT concentrationsare found in the processing site, particularly around structureswhere Hg was released from ore through high-temperatureprocessing equipment. At Puhipuhi, the most contaminatedareas sampled were those historically exposed to Hg vapours(exhaust fan, condensing towers and cooling troughs). GaseousHg0 levels detected in structures and topsoil were quite lowcompared to other Hg mine sites where HgT in soils alsoexceeded 1000 ppm (DeGraff et al. 2007; Rimondi et al. 2012).We speculate that volatilisation of Hg0 from processing areasdoes not contribute significantly to contamination at Puhipuhi.The advection and dispersion of insoluble and soluble Hgspecies across the site from rainfall runoff and deflation oftopsoil could be a major control on Hg contamination atPuhipuhi. This is supported by elevated HgT concentrations insoils from areas not directly involved in the separation andcollection of Hg ore, such as the water holding tanks at Site F. Amore extensive soil survey, including deeper soil profiles, andthe inclusion of more sampling locations removed from theprocessing area, would help identify the advection and disper-sion of Hg from hotspots across the Puhipuhi mine site.

At Puhipuhi, the mine waste tips are located up-gradientfrom the Waikiore Stream. Elevated filtered HgT and MeHgTconcentrations measured in Waikiore Stream near the dumpsites suggest that Hg has been released from waste rock and theprocessing areas, potentially to undergo methylation. Thepositioning and content of the waste rock tips therefore posesa significant contamination risk, as Hg can be released frommine waste calcine and transported into the local catchment.Future investigations should further assess the potential for Hgdispersion by Waikiore Stream from the site into the receivingenvironment.

ReferencesANZECC, ARMCANZ 2000. Australian and New Zealand guidelines

for fresh and marine water quality. Australian and New ZealandEnvironment and Conservation Council and Agriculture andResource Management Council of Australia and New Zealand,Canberra. Pp. 1–103.

ARC 2001. Background concentrations of inorganic elements in soilsfrom the Auckland region. ARC Technical Publication. http://www.aucklandcity.govt.nz/council/documents/ (accessed 29 July 2013).

Baldi F 1988. Mercury pollution in the soil and mosses around ageothermal plant. Water, Air, and Soil Pollution 38: 111–119.

Benoit JM, Gilmour CC, Heyes A, Mason RP, Miller CL 2003.Geochemical and biological controls over methylmercury produc-tion and degradation in aquatic ecosystems. ACS SymposiumSeries 835: 262–297.

Benoit JM, Gilmour CC 1999. Sulfide controls on mercury speciationand bioavailability to methylating bacteria in sediment. Environ-mental Science and Technology 33: 951–951.

Butcher M 2010. The Puhipuhi Mercury Mine history and sitedescription. http://www.doc.govt.nz/conservation/historic/by-region/northland/whangarei/puhipuhi-mercury-mines (accessed 26April 2012).

Christie T, Brathwaite B 1995. Mineral commodity report 8—Mercury.New Zealand Mining 17: 34–39.

Chrystall L, Rumsby A 2009. Mercury inventory for New Zealand2008. Technical Report to the New Zealand Ministry of Environ-ment. Auckland, Pattle Demalore Partners Ltd. 136 p.

Compeau GC, Bartha R 1985. Sulfate-reducing bacteria: principalmethylators of mercury in anoxic estuarine sediment. Applied andEnvironmental Microbiology 50: 498–502.

Craw D, Chappell D, Black A 2002. Surface run‐off from mineralisedroad aggregate, Puhipuhi, Northland, New Zealand. New ZealandJournal of Marine and Freshwater Research 36: 105–116.

Craw D, Chappell D, Reay A 2000. Environmental mercury andarsenic sources in fossil hydrothermal systems, Northland, NewZealand. Environmental Geology 39: 875–887.

Davey HA, van Moort JC 1986. Current mercury deposition at NgawhaSprings, New Zealand. Applied Geochemistry 1: 75–93.

DeGraff JV, Rogow M, Trainor P 2007. Approaches to contaminationat mercury mill sites: Examples from California and Idaho.Reviews in Engineering Geology 17: 115–134.

Eller PM, Cassinelli ME 1994. NIOSH manual of analytical methods.Cincinnati, OH, National Institute for Occupational Safety andHealth.

Ferrar HT 1925. The geology of the Whangarei–Bay of Islandssubdivision, Kaipara division. By authority WAG Skinner.Wellington, Government Printer.

Ganguli PM 2000. Mercury speciation in drainage from the New Idriamercury mine, California. Environmental Science and Technology34: 47–73.

Gilmour CC, Riedel GS, Ederington MC, Bell JT, Gill GA, Stordal MC1998. Methylmercury concentrations and production rates across atrophic gradient in the northern Everglades. Biogeochemistry 40:327–345.

Gosar M, Teršič T 2012. Environmental geochemistry studies in thearea of Idrija mercury mine, Slovenia. Environ Geochem Health34(1): 27–41.

Gray JE, Theodorakos PM, Bailey EA, Turner RR 2000. Distribution,speciation, and transport of mercury in stream-sediment, stream-water, and fish collected near abandoned mercury mines insouthwestern Alaska, USA. The Science of the Total Environment260(1–3): 21–33.

Gray JE, Crock JG, Fey DL 2002a. Environmental geochemistry ofabandoned mercury mines in West-Central Nevada, USA. AppliedGeochemistry 17: 1069–1079.

Gray JE, Crock JG, Lasorsa BK 2002b. Mercury methylation atmercury mines in the Humboldt River Basin, Nevada, USA.Geochemistry: Exploration, Environment, Analysis 2: 143–149.

Gray JE 2003. Leaching, transport, and methylation of mercury in andaround abandoned mercury mines in the Humboldt River Basin

Mercury contamination in New Zealand 9

Dow

nloa

ded

by [

128.

250.

4.11

2] a

t 15:

09 2

9 Ja

nuar

y 20

15

and surrounding areas, Nevada. US Geological Survey Bul-letin 15.

Gray JE, Greaves IA, Bustos DM, Krabbenhoft DP 2003. Mercury andmethylmercury contents in mine-waste calcine, water, and sedi-ment collected from the Palawan Quicksilver Mine, Philippines.Environmental Geology 43(3): 298–307.

Gray JE, Plumlee GS, Morman SA, Higueras PL, Crock JG, LowersHA et al. 2010. In vitro studies evaluating leaching of mercuryfrom mine waste calcine using simulated human body fluids.Environmental Science and Technology 44: 4782–4788.

Henderson J 1944. Cinnabar at Puhipuhi and Ngawha, North Auck-land. NZ Journal of Science and Technology 26: 47–60.

Hoggins FE, Brooks RR 1973. Natural dispersion of mercury fromPuhipuhi, Northland, New Zealand. New Zealand Journal ofMarine and Freshwater Research 7: 125–132.

Kocman D, Kanduč T, Ogrinc N, Horvat M 2010. Distribution andpartitioning of mercury in a river catchment impacted by formermercury mining activity. Biogeochemistry 104: 183–201.

Krabbenhoft DP, Engstrom D, Gilmour C, Harris R, Hurley JP andMason RP 2007. Monitoring and evaluating trends in sedimentand water indicators. In: Harris R, Krabbenhoft DP, Mason R,Murray MW, Reash R, and Saltman T eds. Ecosystem responsesto mercury contamination: indicators of change. Webster, NY:SETAC. Pp. 47–86.

Linstrom PJ, Mallard WG 2001. NIST Chemistry webbook. NISTstandard reference database No. 69.

Lodenius M 1981. Regional distribution of mercury in Hypogymniaphysodes in Finland [air pollution]. Ambio 10.

McLean JE, Bledsoe BE 1992. Behaviour of metals in soils (EPAGround Water Issue, EPA 540-S-92-018: 25 p.). Washington, DC:Environmental Protection Agency.

Ministry of Health 2013. Guidelines for drinking-water qualitymanagement for New Zealand 2013. 3rd edition. Wellington,Ministry of Health.

Navarro A, Biester H, Mendoza JL, Cardellach E 2005. Mercury specia-tion and mobilization in contaminated soils of the Valle del AzogueHg mine (SE, Spain). Environmental Geology 49: 1089–1101.

Northland Regional Council 2002. Groundwater. In: NorthlandRegional Council State of the Environment Report 2002.Pp. 115–137.

O’Brien J 1946. Mines Statement. Appendix to the Journals of theHouse of Representatives Pp 1–56.

Olson ML, DeWild JF 1999. Techniques for the collection and species-specific analysis of low levels of mercury in water, sediment, andbiota. US Geological Survey Water-Resources InvestigationsReport. Pp. 99–4018.

Olund SD 2004. Methods for the preparation and analysis of solids andsuspended solids for total mercury. Reston, VA: US Department ofthe Interior, US Geological Survey.

Qiu G, Feng X, Wang S, Shang L 2005. Mercury and methylmercuryin riparian soil, sediments, mine-waste calcines, and moss fromabandoned Hg mines in east Guizhou province, southwesternChina. Applied Geochemistry 20: 627–638.

Rimondi V, Gray JE, Costagliola P, Vaselli O, Lattanzi P 2012.Concentration, distribution, and translocation of mercury and

methylmercury in mine-waste, sediment, soil, water, and fishcollected near the Abbadia San Salvatore mercury mine, MonteAmiata district, Italy. The Science of the Total Environment 414:318–327.

Rytuba JJ 2000. Mercury mine drainage and processes that control itsenvironmental impact. The Science of the Total Environment 260:57–71.

Rytuba JJ 2003. Mercury from mineral deposits and potential environ-mental impact. Environmental Geology 43(3): 326–338.

Scudder BC, Chasar LC, Wentz DA, Bauch NJ, Brigham ME, MoranPW et al. 2009. Mercury in fish, bed sediment, and water fromstreams across the United States, 1998–2005. Scientific Investiga-tions Report 2009–5109.

Siegel FR ed. 2002. Environmental geochemistry of potentially toxicmetals. Berlin, Springer. 218 p.

Steinnes E 2013. Mercury. In: Alloway BJ ed. Heavy metals in soils:trace metals and metalloids in soils and their bioavailability.3rd ed. Berlin, Springer. Pp. 411–428.

Stetson SJ, Gray JE, Wanty RB, Macalady DL 2009. Isotopicvariability of mercury in ore, mine-waste calcine, and leachatesof mine-waste calcine from areas mined for mercury. Environ-mental Science and Technology 43: 7331–7336.

Takaya M, Joeng JY, Ishihara N, Serita F, Kohyama N 2006. Fieldevaluation of mercury vapor analytical methods: comparison ofthe “double amalgam method” and ISO 17733. Industrial Health44: 287–287.

Ullrich SM, Tanton TW, Abdrashitova SA 2001. Mercury in theaquatic environment: a review of factors affecting methylation.Critical Reviews in Environmental Science and Technology 31:241–293.

UNEP 2013. Minamata convention on mercury—text and annexes.United Nations Environment Programme. Pp. 1–69.

USEPA 2001. Mercury update: impact on fish advisories. U.S.Environmental Protection Agency Fact Sheet. Pp. 1–10.

USEPA 2013. Mercury—Law and Regulations, U.S. EnvironmentalProtection Agency. http://www.epa.gov/hg/regs.htm (accessed 11September 2012).

Vaselli O, Higueras P, Nisi B, María Esbrí J, Cabassi J, Martínez-Coronado A et al. 2013. Distribution of gaseous Hg in theMercury mining district of Mt. Amiata (Central Italy): Ageochemical survey prior the reclamation project. EnvironmentalResearch 125: 179–187.

Webb PC 1940. Mines Statement. Appendix to the Journals of theHouse of Representatives. Pp. 1–102.

Webb PC 1941. Mines Statement. Appendix to the Journals of theHouse of Representatives. Pp. 1–80.

Wiederhold JG, Smith RS, Siebner H, Jew AD, Brown GE, Bourdon Bet al. 2013. Mercury isotope signatures as tracers for Hg cycling atthe New Idria Hg mine. Environmental Science and Technology47: 6137–6145.

Yeaple DS 1972. Mercury in bryophytes (moss). Nature 235: 229–230.Yin R, Feng X, Wang J, Bao Z, Yu B, Chen J 2013. Mercury isotope

variations between bioavailable mercury fractions and totalmercury in mercury contaminated soil in Wanshan MercuryMine, SW China. Chemical Geology 336: 80–86.

10 CM Gionfriddo et al.

Dow

nloa

ded

by [

128.

250.

4.11

2] a

t 15:

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